EPA/600/R-10/100 | August 2010 | www.epa.gov/nrmrl
United States
Environmental Protection
Agency
Nutrient Control Design Manual
Office of Research and Development
National Risk Management Research Laboratory - Water Supply and Water Resources Division
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EPA/600/R-10/100
August 2010
Nutrient Control Design Manual
by
The Cadmus Group, Inc
57 Water Street
Watertown, MA 02472
Scientific, Technical, Research, Engineering, and Modeling Support (STREAMS)
Task Order 68
Contract No. EP-C-05-058
George T. Moore, Task Order Manager
United States Environmental Protection Agency
Office of Research and Development / National Risk Management Research Laboratory
26 West Martin Luther King Drive, Mail Code 445
Cincinnati, Ohio, 45268
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Foreword
The U.S. Environmental Protection Agency (EPA) is charged by Congress with protecting the Nation's
land, air, and water resources. Under a mandate of national environmental laws, the Agency strives to
formulate and implement actions leading to a compatible balance between human activities and the
ability of natural systems to support and nurture life. To meet this mandate, EPA's research program is
providing data and technical support for solving environmental problems today and building a science
knowledge base necessary to manage our ecological resources wisely, understand how pollutants affect
our health, and prevent or reduce environmental risks in the future.
The National Risk Management Research Laboratory (NRMRL) is the Agency's center for investigation of
technological and management approaches for preventing and reducing risks from pollution that
threaten human health and the environment. The focus of the laboratory's research program is on
methods and their cost-effectiveness for prevention and control of pollution to air, land water and
subsurface resources; protection of water quality in public water systems; remediation of contaminated
sites, sediments and ground water; prevention and control of indoor air pollution; and restoration of
ecosystems. NRMRL collaborates with both public and private sector partners to foster technologies
that reduce the cost of compliance and to anticipate emerging problems. NRMRL's research solutions to
environmental problems by: developing and promoting technologies that protect and improve the
environment; advancing scientific and engineering information to support regulatory and policy
decisions; and providing the technological support and information transfer to insure implementation of
environmental regulations and strategies at the national, state, and community levels.
This publication has been produced as part of the Laboratory's strategic long-term research plan. It is
published and made available by EPA's Office of Research and Development to assist the user
community and to link researchers with their clients.
Sally Gutierrez, Director
National Risk Management Research Laboratory
Nutrient Control Design Manual iii August 2010
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Notice
This document was prepared by The Cadmus Group, Inc. (Cadmus) under EPA Contract No. EP-C-05-058,
Task Order 68. The Cadmus Team was lead by Patricia Hertzler and Laura Dufresne with Senior Advisors
Clifford Randall, Emeritus Professor of Civil and Environmental Engineering at Virginia Tech and Director
of the Occoquan Watershed Monitoring Program; James Barnard, Global Practice and Technology
Leader at Black & Veatch; David Stensel, Professor of Civil and Environmental Engineering at the
University of Washington; and Jeanette Brown, Executive Director of the Stamford Water Pollution
Control Authority and Adjunct Professor of Environmental Engineering at Manhattan College.
Disclaimer
The views expressed in this document are those of the individual authors and do not necessarily, reflect
the views and policies of the U.S. Environmental Protection Agency (EPA). Mention of trade names or
commercial products does not constitute endorsement or recommendation for use. This document has
been reviewed in accordance with EPA's peer and administrative review policies and approved for
publication.
Nutrient Control Design Manual iv August 2010
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Abstract
The purpose of this EPA design manual is to provide updated, state-of-the-technology design
guidance on nitrogen and phosphorus control at municipal Wastewater Treatment Plants (WWTPs).
Similar to previous EPA manuals, this manual contains extensive information on the principles of
biological nutrient removal and chemical phosphorus removal to serve as the basis for design. A detailed
description of technologies, both conventional and emerging, serves as a resource for preliminary
technology selection. Because most WWTPs in the United States are equipped with secondary
treatment, the focus of this design manual is on retrofits to add nutrient removal to existing WWTPs
rather than on new treatment plant design, although guidance for greenfield design is presented. Also
new from previous versions, design guidance herein is based on the use of mathematical models and
simulators. Simulators allow designers to study kinetic- as well as time-based solutions while
determining the total mass balances of many constituents. They have become increasingly powerful,
easy to use, and widely accepted for the design of biological nutrient removal facilities. The manual also
includes new information on emerging issues in the industry such as sustainability in wastewater
treatment design and operation, nutrient recovery and reuse, effluent dissolved organic nitrogen, and
measurement of low phosphorus concentrations.
This report was submitted in fulfillment of EPA Contract No. EP-C-05-058, Task Order 68, by The
Cadmus Group, Inc. under the sponsorship of the United States Environmental Protection Agency. This
report covers a period from November 2007 through April 2010 and represents work completed as of
April 2010.
Nutrient Control Design Manual v August 2010
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Contents
Acronyms and Abbreviations xviii
Acknowledgements xxii
1. Introduction 1-1
1.1 History of Nutrient Removal 1-1
1.2 Purpose and Scope of this Manual 1-3
1.3 Manual Organization 1-5
1.4 References 1-6
2. Need for and Benefits of Nitrogen and Phosphorus Removal 2-1
2.1 Introduction 2-1
2.2 Sources of Nitrogen and Phosphorus in Wastewater 2-2
2.2.1 Nitrogen 2-2
2.2.2 Phosphorus 2-3
2.3 Status of Wastewater Treatment in the United States 2-3
2.4 Nutrient Impairment of U.S. Waterways 2-5
2.4.1 Northern Gulf of Mexico 2-5
2.4.2 Chesapeake Bay 2-6
2.4.3 Great Lakes 2-6
2.4.4 Long Island Sound 2-6
2.5 Climate Change Impacts 2-7
2.6 Federal and State Regulations and Initiatives to Reduce Nutrient Pollution 2-8
2.6.1 Water Quality Standards 2-8
2.6.2 Total Maximum Daily Loads (TMDLs) 2-10
2.6.3 NPDES Permitting 2-10
2.6.4 Water Quality Trading 2-11
2.6.5 Technology Evaluation and Guidance 2-12
2.7 Industry Initiatives—The WERF Nutrient Removal Challenge 2-13
2.8 Benefits of Nutrient Removal 2-14
2.8.1 Improved Plant Performance 2-14
2.8.2 Co-Removal of Emerging Contaminants 2-14
2.8.3 Nutrient Recovery and Reuse 2-15
2.9 Challenges of Nutrient Removal 2-15
2.9.1 Energy Requirements 2-15
2.9.2 Release of Nitrous Oxide 2-17
2.10 References 2-18
3. Principles of Phosphorus Removal by Chemical Addition 3-1
3.1 Introduction 3-1
3.2 Available Forms of Metal Salts and Lime 3-1
3.3 Equations and Stoichiometry 3-2
3.3.1 Removable Phosphorus 3-2
3.3.2 Reactions of Metal Salts and Phosphorus 3-2
3.3.3 Reactions of Lime with Phosphorus 3-5
3.4 Solids Separation Processes 3-5
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3.5 Effects on Sludge Production and Handling 3-6
3.6 Two Factors that May Limit the Ability of Plants to Achieve
Very Low Effluent Levels 3-7
3.7 References 3-8
4. Principles of Biological Nitrogen Removal 4-1
4.1 Introduction 4-1
4.2 Nitrogen Removal by Biomass Synthesis 4-2
4.3 Microbiology of Nitrification 4-3
4.4 Reactions and Stoichiometry of Nitrification 4-5
4.5 Nitrification Kinetics 4-6
4.5.1 AOB kinetics 4-10
4.5.2 NOB kinetics 4-13
4.5.3 Effects of Temperature and Dissolved Oxygen on Nitrification Kinetics....4-14
4.5.4 AOB and NOB Kinetics at High Temperature (SHARON® process) 4-16
4.6 Inhibitory Effects of Environmental Conditions on Nitrification 4-17
4.7 Denitrification Fundamentals 4-20
4.8 Microbiology of Denitrification 4-20
4.9 Metabolism and Stoichiometry of Heterotrophic Denitrification 4-21
4.10 Biological Denitrification Kinetics with Influent Wastewater 4-22
4.11 Denitrification Carbon Sources and Relative Consumption Ratios 4-24
4.12 Denitrification Kinetics of Exogenous Carbon Sources 4-28
4.12.1 Denitrification Kinetics with Methanol 4-28
4.12.2 Alternative Exogenous Substrates and Denitrification Kinetics 4-30
4.12.3 Acclimation Time and Degradative Ability of Denitrifying Bacteria with
Exogenous Substrates 4-31
4.13 Specific Denitrification Rates (SDNR) 4-32
4.14 Simultaneous Nitrification-Denitrification 4-34
4.15 Metabolism and Stoichiometry and Kinetics of ANAMMOX® 4-35
4.16 Impacts on Sludge Production and Handling 4-36
4.17 Effluent Dissolved Organic Nitrogen 4-36
4.18 References 4-39
5. Principles of Biological Phosphorus Removal 5-1
5.1 Overview of the Biological Phosphorus Removal Process 5-1
5.2 Substrate Requirements 5-3
5.3 Sources of Volatile Fatty Acids 5-5
5.3.1 Fermentation in the Collection System 5-6
5.3.2 Anaerobic Fermentation of Primary or Return Activated Sludge 5-7
5.3.3 Commercial Sources 5-10
5.4 Environmental Conditions 5-11
5.4.1 Dissolved Oxygen and Nitrates in the Anaerobic Zone 5-11
5.4.2 Oxygen in the Aerobic Zone 5-11
5.4.3 pH 5-12
5.4.4 Temperature 5-12
5.4.5 Cations 5-13
5.5 Kinetics 5-13
5.5.1 Solids Retention Time (SRT) 5-13
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5.5.2 Hydraulic Retention Time (HRT) 5-14
5.6 Important Design and Operational Considerations 5-14
5.6.1 Avoiding Secondary Release of Phosphorus 5-14
5.6.2 Avoiding Backmixing 5-16
5.6.3 Flow and Load Balancing 5-17
5.7 Impacts on Sludge Processing and Handling 5-17
5.8 References 5-18
6. Overview of Nitrogen and Phosphorus Removal Technologies 6-1
6. 1 Introduction 6-1
6.2 Nitrogen Removal Technologies 6-2
6.2.1 Nitrogen Removal in Single Process Unit 6-3
6.2.1.1 Modified Ludzack-Ettinger (MLE) Process 6-3
6.2.1.2 4-Stage Bardenpho 6-5
6.2.1.3 MLE or 4-Stage Bardepho with Membrane Bioractor (MBR) 6-5
6.2.1.4 Sequencing Batch Reactor (SBR) 6-6
6.2.1.5 Oxidation Ditch with Anoxic Zone 6-7
6.2.1.6 Step Feed Biological Nitrogen Removal (BNR) 6-8
6.2.1.7 Simultaneous Nitrification Denitrification (SNdN) 6-9
6.2.1.8 Integrated Fixed Film Activated Sludge (IFAS) 6-10
6.2.1.9 Moving Bed Biofilm Reactor (MBBR) 6-11
6.2.2 Separate Stage Processes—Nitrification 6-12
6.2.2.1 Suspended Growth Nitrification 6-12
6.2.2.2 Attached Growth Nitrification 6-12
6.2.3 Separate Stage Processes—Denitrification 6-13
6.2.3.1 Denitrification Filters 6-14
6.3 Phosphorus Removal Technologies 6-16
6.3.1 Phosphorus Removal by Chemical Addition 6-16
6.3.2 Biological Phosphorus Removal 6-19
6.3.2.1 Pho-redox (A/O) 6-20
6.3.2.2 Oxidation Ditch with Anaerobic Zone 6-21
6.4 Combined Nitrogen and Phosphorus Removal Technologies 6-22
6.4.1 Biological 6-22
6.4.1.1 3 Stage Pho-redox (A2/0) 6-22
6.4.1.2 5-Stage Bardenpho 6-23
6.4.1.3 University of Capetown (UCT), Modified UCT, and Virginia Initiative
Project (VIP) 6-24
6.4.1.4 Westbank 6-26
6.4.1.5 Oxidation Ditch with Anoxic and Anaerobic Zones 6-26
6.4.1.6 Sequencing Batch Reactor (SBR) 6-27
6.4.2 Hybrid Chemical / Biological 6-27
6.4.2.1 Blue Plains Process 6-28
6.4.2.2 Biological-Chemical Phosphorus and Nitrogen Removal (BCFS)
Process 6-28
6.5 Effluent Filtration 6-29
6.5.1 Conventional Down-flow Filters 6-29
6.5.2 Continuous Backwashing Upflow Sand Filters (Dynasand) 6-29
6.5.3 Pulsed Bed Filters 6-30
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6.5.4 Traveling-Bridge Filters 6-30
6.5.5 Discfilters 6-30
6.5.6 Membrane Filters 6-31
6.6 Sidestream Management 6-32
6.7 Technology Performance 6-34
6.7.1 Removal Efficiencies of BNR Technologies - General Discussion 6-36
6.7.2 Technology Performance Statistics based on Full-Scale Operating Data... 6-37
6.8 Factors in Simultaneously Achieving Low Nitrogen and Phosphorus Effluent
Concentrations 6-42
6.9 References 6-43
7. Establishing Design Objectives 7-1
7.1 Introduction 7-1
7.2 Characterizing Existing Treatment 7-2
7.3 Design Flow Rates 7-3
7.3.1 Characterizing Existing Flow 7-3
7.3.2 Projecting Future Conditions 7-5
7.3.3 Setting Design Flow Rates 7-6
7.4 Wastewater characteristics 7-7
7.4.1 Data Collection 7-7
7.4.2 Data Verification 7-10
7.5 Target Effluent Concentrations for Total Nitrogen and Total Phosphorus 7-15
7.6 Goals for Reliability, Sustainability, and Process Flexibility 7-16
7.7 Sludge Treatment Options 7-18
7.8 Site Constraints 7-18
7.9 Selecting an Overall Process Design Factor 7-19
7.10 References 7-20
8. Selecting Candidate Treatment Processes for Plant Upgrades 8-1
8.1 Introduction 8-1
8.2 Technology Selection Factors 8-1
8.2.1 Seasonal Permit Limits 8-1
8.2.2 Footprint 8-2
8.2.3 Hydraulic Considerations 8-3
8.2.4 Chemical needs 8-3
8.2.5 Available Sludge Treatment and Options 8-3
8.2.6 Energy Considerations 8-4
8.2.7 Staffing and Training Requirements 8-5
8.2.8 Technology Selection Considerations for Small Flow Systems 8-5
8.3 Advantages and Disadvantages of Technology Types 8-6
8.4 Overview of Recommended Approach 8-8
8.5 Recommended Use of Advanced Tools 8-11
8.6 Patent issues 8-12
8.7 References 8-13
9. Design Approach for Phosphorus Removal by Chemical Addition 9-1
9.1 Introduction 9-1
9.2 Selecting a Chemical Precipitant 9-1
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9.2.1 Advantages and Disadvantages of Metal Salts 9-1
9.2.2 Advantages and Disadvantages of Lime 9-3
9.2.3 Costs 9-3
9.3 Selecting Point(s) of Application 9-4
9.4 Determining the Chemical Dose 9-7
9.5 Designing a Chemical Feed System 9-10
9.5.1 Liquid feed systems 9-11
9.5.1.1 Storage tanks 9-11
9.5.1.2 Feed Methods 9-11
9.5.2 Dry Feed Systems 9-12
9.5.2.1 Storage 9-12
9.5.2.2 Feed Methods 9-13
9.5.2.3 Lime Slaking 9-14
9.6 Designing for Rapid Mix and Flocculation 9-15
9.6.1 Types of Mixers 9-15
9.6.2 Design Factors 9-17
9.6.2.1 Velocity Gradient 9-17
9.6.2.2 Power Requirements 9-18
9.6.2.3 Hydraulic Retention Time 9-19
9.6.2.4 Vessel Geometry 9-20
9.6.3 Summary of Typical Design Parameters 9-20
9.7 Solids Separation Processes 9-21
9.7.1 Primary and Secondary Clarification 9-22
9.7.2 Tertiary Processes 9-22
9.8 Operational Factors 9-22
9.8.1 Dose Control 9-22
9.8.2 Make-up Water 9-23
9.8.3 Sludge Production and Handling 9-23
9.8.4 pH Adjustment 9-25
9.8.5 Effect on Biosolids Applications 9-25
9.9 References 9-25
10. Design Approach for Biological Nutrient Removal 10-1
10.1 Introduction 10-1
10.2 Preliminary Design Approach 10-3
10.3 Overview of Recommended Approach for Plant Modeling 10-5
10.4 Establishing Modeling Objectives and Requirements 10-7
10.4.1 Intended Model Use 10-7
10.4.2 Goals for Model Accuracy 10-7
10.4.3 Dynamic vs. Steady State Simulation 10-8
10.5 Selecting a Process Simulation Model 10-9
10.6 Data Collection 10-12
10.6.1 Process Configuration 10-13
10.6.2 Operating Conditions 10-16
10.7 Characterization of Organic Material 10-16
10.7.1 Relationship of Organic Material and Suspended Solids in Wastewater... 10-20
10.7.2 Methods for Determining COD Fractions 10-22
10.7.3 Data Checks 10-25
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10.8 Characterization of Nutrient Fractions 10-26
10.8.1 Nitrogen 10-26
10.8.2 Phosphorus 10-29
10.9 Kinetic and Stoichiometric Parameters 10-32
10.10 Calibration 10-33
10.11 Validation 10-37
10.12 Simulation of Design Alternatives for Nutrient Removal 10-38
10.13 Additional Procedures for Design 10-39
10.13.1 Sequencing Batch Reactors (SBRs) 10-39
10.13.2 Denitrification Filters 10-40
10.13.3 Primary Sludge Fermenters 10-41
10.14 Design Checks for Biological Nitrogen and Phosphorus Removal 10-42
10.15 References 10-47
11. Design Approach for Effluent Filtration 11-1
11.1 Introduction 11-1
11.2 Selection of Filtration Technology 11-2
11.3 Granular Media Filters 11-3
11.3.1 Influent Water Quality 11-4
11.3.2 Media Specifications 11-4
11.3.3 Filter Loading Rate 11-6
11.3.4 Headless 11-6
11.3.5 Backwash Requirements 11-7
11.3.6 Flow Control 11-9
11.4 Cloth Media Filters 11-10
11.5 Low-Pressure Membranes 11-11
11.5.1 Membrane Material 11-12
11.5.2 Membrane Configuration 11-13
11.5.3 Process Considerations 11-14
11.5.4 Pressure Drop 11-15
11.5.5 Flux Determination 11-15
11.5.6 Performance Data 11-16
11.6 Emerging Filtration Technologies for Phosphorus Removal 11-16
11.6.1 Two-Stage Filtration 11-16
11.6.2 Iron Oxide Coated Media 11-17
11.7 References 11-18
12. Operation and Optimization to Enhance Nutrient Removal 12-1
12.1 Introduction 12-1
12.2 Analysis of Existing Operations 12-1
12.2.1 Data Analysis 12-2
12.2.2 Use of Process Simulation Models 12-5
12.3 Incorporating SCADA and other Instrumentation 12-6
12.4 Common Operational Changes 12-6
12.4.1 Adjust SRT 12-6
12.4.2 Adjust Aeration Rates 12-7
12.4.3 Add Baffles to Create High Food to Microorganism (F/M) Conditions 12-7
12.4.4 Change Aeration Settings in Plug Flow Basins 12-7
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12.4.5 Minimize Impact of Recycle Streams 12-8
12.4.6 Reconfigure Flow through Existing Units 12-8
12.4.7 Increase VFAs for Biological Phosphorus Removal 12-9
12.5 References 12-10
13. Instrumentation and Controls 13-1
13.1 Introduction 13-1
13.2 Factors in Selecting Instrumentation 13-2
13.3 Basic Online Instrumentation 13-3
13.3.1 Flow 13-3
13.3.2 Total Suspended Solids (TSS) 13-4
13.3.3 Sludge Blanket Depth 13-5
13.3.4 Dissolved Oxygen (DO) 13-5
13.3.5 pH 13-5
13.3.6 ORP 13-6
13.4 Online Instrumentation for Nutrient Control 13-6
13.4.1 Nitrogen Compounds 13-6
13.4.2 Phosphate and Total Phosphorus 13-7
13.4.3 NADH (active biomass) 13-8
13.4.3 Respirometry 13-8
13.5 Types of Control 13-9
13.5.1 Feed-forward 13-9
13.5.2 Feedback 13-10
13.5.3 Feed-forward and feedback 13-10
13.5.4 Cascade 13-10
13.5.5 Advanced Control 13-10
13.6 Control Equipment—SCADA 13-11
13.7 References 13-13
14. Sustainable Nutrient Recovery and Reuse 14-1
14.1 Introduction 14-1
14.2 Separating and Treating Waste On-Site 14-1
14.3 Using Wastewater Treatment Byproducts 14-2
14.3.1 Durham, OR, Advanced Wastewater Treatment Facility 14-3
14.3.2 East Bay Municipal Utility District, CA 14-4
14.4 References 14-5
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Appendices
Appendix A. Recommendations for Methanol Safety
Appendix B. Organic Compounds and Inhibitory Concentrations to Nitrification
Appendix C. Mathematical Models for Wastewater Treatment
Nutrient Control Design Manual xiii August 2010
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Tables
Table 3-1. Chemical Precipitants 3-2
Table 4-1. Phylogeny of Ammonia-Oxidizing Bacteria 4-4
Table 4-2. Phylogeny of Nitrite-Oxidizing Bacteria 4-5
Table 4-3. Summary of Test Results on Measuring Specific Endogenous Decay Coefficient
Rates (All Rates at 20°C) 4-11
Table 4-4. Summary of AOB Nitrification Kinetic Coefficient Values 4-12
Table 4-5. Comparison of Nitrification Half-Velocity Coefficients (mg/L) in MBR and
Conventional Activated Sludge (CAS) Systems 4-12
Table 4-6. Summary of NOB Nitrification Kinetic Coefficient Values 4-14
Table 4-7. NH4-N and NO2-N Concentrations that May Inhibit Nitrification as a Function
of pH at 20°C 4-19
Table 4-8. Heterotrophic Bacteria Kinetic Coefficients in Anoxic/Aerobic Activated
Sludge 4-24
Table 4-9. Biomass Yields Reported for Exogenous Carbon Sources 4-28
Table 4-10. Reported Maximum Specific Growth Rates at 20°C and Temperature
Coefficients for Methanol Utilization under Anoxic and Aerobic conditions 4-29
Table 4-11. Reported Ks values for NO3-N reduction with methanol at 20°C 4-29
Table 4-12. Comparison of Maximum Specific Growth rates for Methanol, Acetate, and Corn
Syrup at High and Low Temperatures 4-30
Table 4-13. For BNR Activated Sludge, Ratio of Denitrification Rate with Substrate Addition
to Denitrification Rate with No Addition 4-31
Table 4-14. Ratio of Denitrification Rates for Other Substrates at Day 50 with Ethanol or
Methanol Addition Versus no Addition 4-32
Table 4-15. Range of reported SDNR values in BNR activated sludge treatment 4-33
Table 4.16. ANAMMOX® Bacteria Biokinetic Parameters at 30°C 4-35
Table 5-1. Volatile Fatty Acids Typically Found in Fermented Wastewater 5-4
Table 5-2. Minimum Ratios for Achieving Total Phosphorus Effluent Concentration of
less than 1.0 mg/L 5-4
Table 5-3. Effect of Corrosion and Odor Control Techniques on VFA Production in
Wastewater Collection Systems 5-7
Table 6-1. Matrix of Biological Nitrogen Removal Technologies 6-3
Table 6-2. IFAS Media Types, Applications, and Design Considerations 6-10
Table 6-3. Matrix of Phosphorus Removal Technologies 6-16
Table 6-4. Matrix of Combined Biological Phosphorus and Nitrogen Removal Technologies . 6-22
Table 6-5. Matrix of Tertiary Filtration Technologies 6-29
Table 7-1. Influent Flow Components 7-4
Table 7-2. Flow Characterization 7-4
Table 7-3. Comparison of Flow Rates and Flush Volumes Before and After U.S. Energy
Policy Act 7-5
Table 7-4. Example Permit Limits for Nutrients 7-16
Table 8-1. Advantages and Disadvantages of Technology Types 8-7
Table 8-2. External Carbon Sources 8-12
Table 8-3. IFAS Media Types, Applications, and Design Considerations 8-15
Table 9-1. Advantages and Disadvantages of Common Aluminum and Iron Salts 9-2
Table 9-2. Advantages and Disadvantages of Metal Salt Application Points 9-6
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Table 9-3. Types of Chemical Feeders 9-10
Table 9-4. Commonly Used Equipment for Rapid Mixing 9-16
Table 9-5. Values of NP and NQfor Various Types of Impellers 9-18
Table 9-6. Typical Design Parameters for Turbine and Propeller Mixer 9-21
Table 10-1. Commonly Used Process Simulators 10-10
Table 10-2. COD and Particulate Fractions in Municipal Wastewater 10-19
Table 10-3. TKN Fractions in Municipal Wastewater 10-28
Table 10-4. Total Phosphorus Fractions in Municipal Wastewater 10-31
Table 10-5. Design Checks for Biological Nitrogen Removal 10-43
Table 10-6. Design Checks for Biological Phosphorus Removal 10-45
Table 11-1. Common Filter Media and Characteristics 11-5
Table 11-2. Filter Media Depths and Particle Sizes 11-5
Table 11-3. Membrane Characteristics 11-14
Table 11-4. Advantages and Disadvantages of Membrane Materials 11-15
Table 11-5. Phosphorus Removal Reported From Membrane Pilot Studies 11-19
Table 11-6. Pilot Test Results for the Blue Water Blue PRO® System 11-21
Table 12-1. Recommended Parameters for Data Evaluation 12-3
Table 13-1. Summary of Basic On-Line Instrumentation 13-2
Table 13-2. Comparison of Online Nitrate Analyzers 13-7
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Figures
Figure 2-1. Population served by POTWs nationwide for select years between 1940 and 2004
and projected to 2024 (if all needs are met), organized by wastewater
treatment type 2-5
Figure 2-2. Water quality-based approach of the Clean Water Act 2-10
Figure 4-1. Percent nitrogen removal due to biomass synthesis as a function of SRT and
influent BOD/N ratio 4-3
Figure 4-2. Effect of SRT and temperature on effluent NH4+-N and NO2"-N concentrations
using kinetic data in Table 4-6 and 4-4 for CMAS 4-15
Figure 4-3. Effect of DO concentration on effluent NH4+-N and NO2"-N concentrations
using kinetic data in Table 4-4 and 4-6 and at 15°C for CMAS 4-15
Figure 4-4. Effect of temperature on minimal washout SRT AOB, H and NOB, H from
Hellinga et al. (1998) and AOB, M and NOB, M from Tables 4-2 and 4-4 4-17
Figure 4-5. Ratio of COD required to NO3-N completely reduce NO3-N (CRNO3) as a
function of the biomass yield 4-26
Figure 5-1. Theory of BPR in activated sludge 5-1
Figure 5-2. BPR at a WWTP 5-3
Figure 5-3 Biological Pathways of Methane Formation 5.8
Figure 5-4. Example of secondary release in second anoxic zone 5-14
Figure 6-1. Modified Ludzack-Ettinger (MLE) process 6-4
Figure 6-2. 4-stage Bardenpho process 6-5
Figure 6-3. Common configuration for a membrane bioreactor 4-stage bardenpho treatment
system 6-6
Figure 6-4. Operating periods of a sequencing batch reactor 6-7
Figure 6-5. Example oxidation ditch configuration 6-8
Figure 6-6. Step feed biological nitrogen removal 6-7
Figure 6-7. Downflow denitrification filter 6-14
Figure 6-8. Continuous backwash upflow sand (CBUS) filters 6-15
Figure 6-9. Close-up of continuous backwash upflow sand (CBUS) filter 6-15
Figure 6-10. Densadeg® high rate clarification process flow diagram 6-18
Figure 6-11. CoMagTM process flow diagram 6-19
Figure 6-12. Pho-redox process (A/O) 6-21
Figure 6-13. Oxidation ditch with anaerobic zone 6-22
Figure 6-14. 3 Stage Pho-redox process (A2/O) 6-23
Figure 6-15. 5-stage Bardenpho process 6-24
Figure 6-16. UCT and Modified UCT process 6-25
Figure 6-17. Westbank process 6-26
Figure 6-18. VT2 process schematic 6-27
Figure 6-19. The Blue Plains process 6-28
Figure 6-20. Probability plot of secondary effluent phosphorus data 6-39
Figure 6-21. Technology performance statistics for nitrogen removal plants 6-40
Figure 6-22. Technology performance statistics for phosphorus removal plants 6-40
Figure 7-1 Net sludge production versus solids retention time and temperature 7-14
Figure 9-1. Possible application points for chemical addition (C) 9-4
Figure 9-2. Schematic of common jar testing apparatus 9-9
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Figure 9-3. Typical dry chemical feed system 9-13
Figure 10-1. Unified protocol for activated sludge monitoring 10-6
Figure 10-2. Essential requirements for wastewater treatment process simulation 10-13
Figure 10-3. Example simulator configuration for a biological nutrient removal plant 10-14
Figure 10-4. COD components for municipal wastewater 10-18
Figure 10-5. Relationship between BOD, COD, TSS, and VSS 10-21
Figure 10-6. TKN components for municipal wastewater 10-27
Figure 10-7. Phosphorus components in municipal wastewater 10-30
Figure 11-1. Upflow continuous backflow filter 11-11
Figure 11-2. Operational models of the Fuzzy Filter® 11-12
Figure 11-3. Cutaway view of AquaDisk® cloth media filter 11-13
Figure 11-4. Hollow fiber membrane configuration with inside to outside flow 11-16
Figure 11-5. Parkson Dynasand D2 advanced filter system 11-20
Figure 11-6. Blue-PRO® process 11-21
Figure 12-1. Spatial and temporal profiles of ammonia 12-5
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Acronyms and Abbreviations
A/O
A2/O
AMD
ANAMMOX
AOB
AS
ASCE
ASM
AT3
BABE
BAF
BAR
BCFS
bDON
BHRC
BNR
BOD
BOD5
BPR
CCF
CFD
CIP
CMAS
C/N
COD
COV
CR
CSO
CSTR
CWA
CWSRF
DAF
DO
DON
DSS
EBPR
EDC
EDTA
ENR
EPA
FFS
F/M
FWPCA
FWS
GAO
Anaerobic/Oxic, Pho-redox
Anaerobic/Anoxic/Oxic, 3 Stage Pho-redox
Ammonia Monooxygenase
Anaerobic Ammonia Oxidation
Ammonia Oxidizing Bacteria
Activated Sludge
American Society of Civil Engineers
Activated Sludge Model
Aeration Tank 3
Bio-Augmentation Batch Enhanced
Biological Aerated Filter
Bio-Augmentation Regeneration/Reaeration
Biological Chemical Phosphorus and Nitrogen Removal
Biodegradable Fraction of Dissolved Organic Nitrogen
Ballasted High Rate Clarification Processes
Biological Nutrient Removal
Biochemical Oxygen Demand
Biochemical Oxygen Demand (5-day)
Biological Phosphorus Removal
Continuous Contact Filtration
Computational Fluid Dynamic
Clean in Place
Completely-Mixed Activated Sludge
Carbon to Nitrogen Ratio
Chemical Oxygen Demand
Coefficient of Variation
Consumptive Ratio
Combined Sewer Overflow
Continuous Stirred Tank Reactors
Clean Water Act
Clean Water State Revolving Fund
Dissolved Air Flotation
Dissolved Oxygen
Dissolved Organic Nitrogen
Designated Suspended Solids
Enhanced Biological Phosphorus Removal
Endocrine Disrupting Chemicals
Ethylene Diamine Tetraacetic Acid
Enhanced Nutrient Removal
U.S. Environmental Protection Agency
Fixed-film Systems
Food to Microorganism ratio
Federal Water Pollution Control Act
Free Water Surface
Glycogen Accumulating Organism
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GMP
HRSD
HRT
iDON
I FAS
ISF
ISS
IWA
JHB
LOT
MAUREEN
MBBR
MBR
MCL
MF
MGD
mg/L
MLE
MLSS
MLVSS
MMDF
MUCT
MWRDGC
N
NF
NTU
NOAA
NOB
NPDES
NTT
ORD
ORP
OSHA
OUR
OWASA
OWM
P
PACI
PAH
PAD
PHA
PHB
PHV
PID
PLC
POTW
PPCPs
RAS
Good Modeling Practices
Hampton Roads Sanitation District
Hydraulic Retention Time
Inert Dissolved Organic Nitrogen
Integrated Fixed-Film Activated Sludge
Intermittent Sand Filter
Inert Suspended Solids
International Water Association
Johannesburg Process
Limit of Technology
Mainstream Autotrophic Recycle Enhanced N-removal
Moving-Bed Biofilm Reactor
Membrane Bioreactor
Maximum Contaminant Level
Microfiltration
Million Gallons per Day
Milligrams per liter
Modified Ludzack Ettinger
Mixed Liquor Suspended Solids
Mixed Liquor Volatile Suspended Solids
Maximum Month Design Flow
Modified University of Capetown
Metropolitan Water Reclamation District of Greater Chicago
Nitrogen
Nanofiltration
Nephelometric Turbidity Units
National Oceanic and Atmospheric Administration
Nitrite-Oxidizing Bacteria
National Pollutant Discharge Elimination System
Nitrogen Trading Tool
EPA Office of Research and Development
Oxidation Reduction Potential
Occupational Safety and Health Administration
Oxygen Uptake Rate
Orange Water and Sewer Authority
EPA Office of Wastewater Management
Phosphorus
Polyaluminum Chloride
Polycyclic Aromatic Hydrocarbons
Phosphate Accumulating Organism
Poly-R-hydroxy-alkanoate
Poly-R-hydroxy-butyrate
Poly-hydroxy valerate
Phased Isolation Ditch
Programmable Logic Controller
Publicly Owned Treatment Works
Pharmaceuticals and Personal Care Products
Return Activated Sludge
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RBC
rbCOD
rDON
RO
RSF
SAV
SBCOD
SBR
SCADA
SCM
SDNR
SHARON
SNdN
SRT
SSO
STAC
SWIS
TAL
TAN
TDS
TKN
TMDL
TN
TP
TSS
TUDP
UCT
UF
UOSA
USDA
USEPA
USGS
VFA
VIP
VSS
WAS
WEF
WEFTEC
WERF
WQS
WWTP
Rotating Biological Contactor
Readily Biodegradable Chemical Oxygen Demand
Recalcitrant Dissolved Organic Nitrogen
Reverse Osmosis
Recirculating Sand Filters
Submerged Aquatic Vegetation
Slowly Biodegradable Chemical Oxygen Demand
Sequencing Batch Reactors
Supervisory Control and Data Acquisition
Surface Complexation Modeling
Specific Denitrification Rate
Single Reactor High-Activity Ammonia Removal Over Nitrite
Simultaneous Nitrification-Denitrification
Solids Retention Time
Sanitary Sewer Overflow
Chesapeake Bay Program Scientific and Technical Advisory
Committee
Subsurface Wastewater Infiltration System
Technology Achievable Limit
Total Ammonia Nitrogen
Total Dissolved Solids
Total Kjeldahl Nitrogen
Total Maximum Daily Loads
Total Nitrogen
Total Phosphorus
Total Suspended Solids
Bio-P Model of the Delft University of Technology
University of Capetown Process
Ultrafiltration
Upper Occoquan Sewage Authority
U.S. Department of Agriculture
U.S. Environmental Protection Agency
U.S. Geological Survey
Volatile Fatty Acid
Virginia Initiative Plant
Volatile Suspended Solids
Waste Activated Sludge
Water Environment Federation
Water Environment Federation Technical Exhibition and
Conference
Water Environment Research Foundation
Water Quality Standards
Wastewater Treatment Plant
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Acknowledgements
The principle authors of this document, titled "Nutrient Control Design Manual/' were:
Dr. Clifford Randall, Professor Emeritus of Civil and Environmental Engineering at Virginia Tech and
Director of the Occoquan Watershed Monitoring Program
Dr. James Barnard, Global Practice and Technology Leader at Black & Veatch
Dr. H. David Stensel, Professor of Civil and Environmental Engineering at the University of Washington
Laura Dufresne, Senior Engineer, the Cadmus Group, Inc.
EPA technical direction and oversight were provided by Dan Murray, EPA Office of Research and
Development, National Risk Management Laboratory.
EPA technical reviews of the document were performed by:
EPA Office of Research and Development
Donald Brown
Douglas Grosse
Richard Brenner
James Smith
Marc Mills
Jeffry Yang
Edwin Barth
EPA Headquarters
Donald Anderson
Phil Zahreddine
James Wheeler
EPA Regions
David Pincumbe, Region 1
Roger Janson, Region 1
Russ Martin, Region 5
Dave Ragsdale, Region 10
External technical reviews of the document were performed by
Jeanette Brown, Stamford Water Pollution Control Authority
Tanya Spano, Metropolitan Washington Council of Governments (MWCOG)
S. Joh Kang, Tetra Tech
Nutrient Control Design Manual xxi August 2010
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The following members of the Ohio Water Environment Association:
Dale E. Kocarek, Stantec Consulting, Inc.
William Barhorst, ARCADIS
Dennis P. Meek, DM Engineering
Kim Riddell, City of Delphos
Paul Fletcher, Jones & Henry Engineers
Jason Tincu, City of Xenia
Gary Hickman, City of Columbus
Roger F. Gyger, m2t Technologies
Ted Marten, City of Twinsburg
David Wilson, Butler County Water and Sewer
Karen Harrison, Jordan, Jones, Goulding
Mary Long, Black & Veatch
Robert Hollis, Summit County
Rick Noss, Stantec Consulting
Theping Chen, AECOM
Shaun Beauchesne, Hach Company
David Frank, ARCADIS
The following members of the Water Environment Research Foundation (WERF) Nutrients Challenge
Team:
JB Neethling, HDR Engineering, Inc.
Mario Benisch, HDR Engineering, Inc.
Amit Pramanik, WERF
Diagrams for illustration of specific concepts were provided by:
Dr. Clifford Randall, Virginia Tech
Dr. James Barnard, Black and Veatch
Dr. H. David Stensel, University of Washington
Nutrient Control Design Manual xxii August 2010
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1. Introduction
Chapter 1 covers:
1.1 History of Nutrient Removal
1.2 Purpose and Scope of this Manual
1.3 Manual Organization
1.4 References
1.1 History of Nutrient Removal1
Biological nutrient removal (BNR) at wastewater treatment plants (WWTP) began in the early
1960s. Pioneers such as Ludzack and Ettinger (1961) and Wuhrman (1964) made efforts to develop
biological nitrogen removal (nitrification-denitrification) wastewater treatment systems. Levin and
Shapiro (1965) researched biological phosphorus removal, and developed a patented process for it,
known as PhoStrip. However, the systems devised by Ludzack, Ettinger, and Wuhrman, did not utilize an
internal recycle to obtain significant utilization of the influent biochemical oxygen demand (BOD), and
the proposed biological mechanisms of the PhoStrip process remained controversial because its two
final steps were the release of phosphorus from activated sludge under anaerobic conditions and then
chemical precipitation of the released phosphorus in a separate reactor.
The major process development breakthroughs for biological removal of both nitrogen and
phosphorus utilizing the influent BOD resulted from the work of James Barnard in South Africa in the
early 1970s. He first developed a single sludge process configuration with internal recycle that utilized
the influent BOD for denitrification (1973). It subsequently became the standard nitrogen removal
process for the wastewater industry. It is now known as the modified Ludzack-Ettinger (MLE) process.
He also demonstrated that anaerobic-aerobic sequencing of activated sludge, with influent BOD first
flowing into the anaerobic zone, was necessary to obtain robust biological phosphorus removal (BPR).
This discovery was first published in 1975. Theoretical support that the mechanism was biological and
not chemical was supplied by Fuhs and Chen (1975) in the same year. Barnard developed several
process configurations for both separate and combined biological removal of nitrogen and phosphorus.
A four stage anoxic-aerobic-anoxic-aerobic process designed primarily for nitrogen removal was
patented as the Bardenpho Process (1978). The five stage version, created by adding an anaerobic zone
as the first stage became known as the Modified Bardenpho Process.
Also during the mid-1970s, an anaerobic-aerobic wastewater treatment configuration was being
developed in the United States for control of filamentous growths in activated sludge. This process was
patented by Marshall Spector and acquired by Air Products and Chemical, Inc. They learned from
Barnard that anaerobic-aerobic sequencing of activated sludge also could be used to accomplish BPR
and patented the configuration as the Anaerobic-Oxic (AO) process, which was identical to the Phoredox
configuration developed by Barnard in South Africa. They then combined it with an anoxic zone and
patented the resulting configuration as the Anaerobic-Anoxic-Oxic (A2/O) process, again identical to a
configuration developed by Barnard. At this time, the detrimental impacts of nitrate recycle in return
1 By Dr. Clifford Randall, Professor Emeritus, Virginia Tech
Nutrient Control Design Manual 1-1 August 2010
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activated sludge (RAS) to the anaerobic zone on BPR performance was not fully understood, and many
of the PhoStrip, Phoredox/A2/O, and Modified Bardenpho plants were removing phosphorus erratically.
The South African Government requested that Professor Gerrit Marais and his co-workers at the
University of Cape Town investigate and resolve the issue. They developed a modification of the
Phoredox/A2/O configuration, dubbed the University of Cape Town (UCT) process, that first sent the
RAS to the anoxic zone then added a second internal recycle to recycle denitrified mixed liquor from the
effluent of the anoxic zone back to the influent of the anaerobic zone. Based on the supposition that
denitrification would occur only in the anoxic zone, a modified version of the UCT process was
developed for wastewaters with a high Total Kjeldahl Nitrogen (TKN) to BOD ratio.
BNR was introduced to North America in the early 1980s through implementation of BNR
facilities at Kelowna, BC, Canada, and at Orange County, FL BNR was introduced to the Chesapeake Bay
region in 1984 by a seminar and a workshop organized by Dr. Clifford Randall (Virginia Tech) and held at
Richmond, VA. Then, working with the Hampton Roads Sanitation District (HRSD) and the Virginia Water
Pollution Control Board, a pilot plant study of a high rate UCT process was conducted at the HRSD
Lambert's Point primary treatment plant in 1985-86, and followed by full-scale research-demonstrations
of the A/O, A2/O and UCT processes at the HRSD York River Plant from 1986-90. Overlapping the York
River demonstrations, which resulted in patenting of the Virginia Initiative Plant (VIP) BNR process, were
full-scale demonstrations of BNR (both N and P removal) at the Anne Arundel County, MD, Maryland
City WWTP, and the Bowie, MD, WWTP. Also overlapping these events were the design and
construction of the Mauldin Road WWTP, Greeneville, SC, and modification of two plants in Charlotte,
NC. North American BNR developments moved rapidly in the late 1980s and early 1990s, resulting in
BPR and BNR implementation, design and construction at sites as diverse as the Bonnybrook WWTP,
Alberta, Canada, Hillsborough County, FL, Frederick, MD, Atlanta, GA, and modification of the Howard
County, MD, WWTP from the PhoStrip to the Phoredox/A2/O configuration.
BNR began to be implemented in Europe on a widespread basis in 1987, first in Germany and
The Netherlands, followed by Denmark, Austria, Czech Republic, Italy and France. Schreiber Klarenlagen
with their unique Simultech Process wherein BPR, nitrification, and denitrification all occurred
simultaneously in one continuous flow reactor was a pace setter in Germany while Kruger, Inc., led the
way in Denmark under the guidance of Professor Poul Harremoes and his co-workers at the Danish
Institute of Technology.
The engineering art of BNR has progressed towards maturity during the past two decades with
the addition of advanced practices such as pre-fermentation of primary sludge to enhance BPR,
integration of fixed-film media into activated sludge (IFAS) to enhance nitrogen removal, utilization of
biological filters for nitrogen removal, and widespread use of tertiary filters for dentrification and
chemical phosphorus removal to lower levels. Recent efforts to develop economical methods for the
nutrient removal in sites with limited space for expansion have resulted in the emergence of two
innovative technological approaches:
1) Technologies such as membrane bioreactors or ballasted flocculation to remove suspended
solids to very low concentrations and simultaneously eliminate or greatly reduce the size of
secondary settling basins.
2) Sidestream processes such as SHARON, ANAMMOX, IN-NITRI and others to either remove
nitrogen from ammonia-rich flows from sludge processing or enhance removal in the main
stream process.
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Nitrogen removal has been widely implemented along the Connecticut coast of Long Island
Sound. More stringent effluent standards, typically 3.0 milligrams per liter (mg/L) total nitrogen (TN) and
0.1 or lower mg/L total phosphorus (TP), in regions such as the Chesapeake Bay watershed, coastal
areas of North Carolina, Okanagan Lake area of British Columbia, Canada, mid-Colorado and Kalispell,
WY, have advanced the art from BNR to enhanced nutrient removal (ENR). A combination of BNR,
chemical additions and effluent filtration are typically used to accomplish ENR.
A clear trend of the wastewater treatment industry is a greater emphasis on incorporating
elements of recycle, recovery, and reuse into plant design and operation. Sustainable nutrient recovery
and reuse is gaining national and international attention as wastewater utilities look for ways to
decrease energy costs and greenhouse gas emissions, utilize excess capacity, generate new revenue,
and address ever more stringent regulatory requirements. This evolution in thinking is moving
wastewater treatment to enhanced energy efficiency and changing the role of wastewater treatment
facilities from waste generators to resource providers.
1.2 Purpose and Scope of this Manual
Research and technology development through the mid-70s were the basis for EPA's first design
manual for nitrogen control technologies. This document, "Process Design Manual for Nitrogen
Control," (EPA, 1975) was published in 1975. This manual covered a broad range of processes that were
being evaluated and applied at the time. The intent of the manual was to present design information for
technologies that appeared to have a viable, practical application to nitrogen control. Two broad
categories of treatment processes were addressed. The first group of processes provides for the
conversion of organic and ammonium nitrogen by oxidation to nitrate nitrogen. These processes are
biological and are generally referred to as nitrification. The second group of processes removes nitrogen
from the wastewater. These processes are also biological, using an anoxic denitrification step with
nitrification. Physical/chemical processes were also presented for nitrogen removal, including ion
exchange, ammonia stripping, and breakpoint chlorination. Between the publication of the first nitrogen
control manual and the update of the manual in 1993, the trend in nitrogen control technology was
almost exclusively towards biological processes. Biological processes became proven and well
demonstrated and were most efficiently expanded or upgraded for biological nitrification or total
nitrogen removal. The focus of the 1993 updated document, "Manual - Nitrogen Control," (EPA, 1993)
was on biological/mechanical processes that were finding widespread application for nitrification and
nitrogen removal.
In 1971, EPA published its first phosphorus control design manual. This manual, "Process Design
Manual for Phosphorus Removal," (EPA, 1971) focused on phosphorus removal methods that involve
chemical precipitation. Primarily, the manual focused on the chemical precipitation of phosphorus using
salts of aluminum and iron, and lime. The chemical application points addressed in the manual were
before primary settling, in the aeration tanks, before final settling, or in a tertiary process. In 1976,
"Process Design Manual for Phosphorus Removal" (EPA, 1976) was updated. Specifically, design
guidance for phosphorus removal using mineral addition and lime addition before primary settling was
revised. Also, guidance for chemical storage, chemical feed systems and residuals handling and disposal
was updated. In 1987, EPA published two technical documents that addressed phosphorus control. The
first was an update to the 1976 Process Design Manual for Phosphorus Removal (EPA, 1987a). The
second was a handbook titled, "Handbook - Retrofitting POTWs for Phosphorus Removal in the
Chesapeake Bay Drainage Basin." (EPA, 1987b) The update of the design manual included a major
Nutrient Control Design Manual 1-3 August 2010
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addition of guidance for biological phosphorus removal. Also, the use of lime addition was not covered
in this update due to its loss of popularity in the 80s. The technical guidance provided in the 1987
handbook was focused on the unique phosphorus removal requirements being applied to municipal
wastewater treatment plants in the Chesapeake Bay watershed. Because of the varying levels of
phosphorus control within the watershed, the handbook included an assessment of technologies for
meeting total phosphorus effluent limits of 0.2 mg/L, 0.5 mg/L, 1 mg/L, and 2 mg/L Because some
treatment plants in the Chesapeake Bay watershed need to control both nitrogen and phosphorus, the
handbook included a chapter titled, "Compatibility of Chemical and Biological Phosphorus Removal with
Nitrogen Control."
In 2007, EPA initiated the process to develop updated design guidance for both nitrogen and
phosphorus removal at municipal WWTPs. The first step was an extensive, state-of-the-technology
review of nitrogen and phosphorus control technologies and techniques currently applied and emerging
at municipal wastewater treatment plants. This technology review culminated with the publication of
the "Nutrient Control Design Manual: State of the Technology Review Report" (USEPA 2009) as an
interim document in the development of the updated design manuals.
The purpose of this EPA design manual is to provide updated, state-of-the-technology design
guidance on nitrogen and phosphorus control at municipal WWTPs to wastewater utility owners and
operators, state and EPA permit writers, and environmental engineering professionals. Similar to
previous EPA manuals, this manual includes extensive information on the principles of biological
nutrient removal and chemical phosphorus removal to serve as the basis for design. A detailed
description of technologies, both conventional and emerging, serves as a resource for preliminary
technology selection. The manual presents new information on emerging issues in the industry such as
sustainability in wastewater treatment design and operation, nutrient recovery and reuse, effluent
dissolved organic nitrogen, and measurement of low phosphorus concentrations. Although this manual
provides some examples of proprietary and emerging technologies, EPA recognizes that the industry is
continually evolving and that new technologies not identified in this manual may emerge in the future.
Because the majority of WWTPs in the United States are equipped with secondary biological
treatment, the focus of this design manual is on process and technology modifications/additions for
nutrient removal at existing WWTPs rather than on new treatment plant design, although guidance for
greenfield design is presented. Also new from previous versions, design guidance herein is based on the
use of mathematical models and simulators. Simulators allow designers to study kinetic- as well as time-
based solutions while determining the total mass balances of many constituents. They have become
increasingly powerful, easy to use, widely accepted, and recommended by WEF and ASCE (2010) for the
design of biological nutrient removal facilities. Earlier versions of EPA nutrient control manuals (USEPA
1993; USEPA 1987a; USEPA 1987b) still contain very useful guidance (including examples) on process
design using hand calculations that can be used for very preliminary sizing or checks on simulation
results.
This manual compliments detailed cost data and in-depth facility case studies published in the
Municipal Nutrient Removal Technologies Reference Document (USEPA 2008a) and analysis of emerging
technologies for nutrient removal presented in the Emerging Technologies Report on Wastewater
Treatment (USEPA 2008b). Both documents are available for download from EPA's website at
http://www.epa.gov/OWM/mtb/publications.htm
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1.3 Manual Organization
This design manual has 14 chapters and 3 appendices. It is generally organized with the theory
of nutrient removal presented first followed by a description of nutrient removal technologies; guidance
on establishing design objectives and selecting candidate treatment processes; and design approaches
for chemical phosphorus removal, biological nutrient removal, and effluent filtration. Later chapters
describe operational improvements for enhancing technology performance and guidance on
instrumentation and controls. The last chapter, Chapter 14, discusses sustainable recovery and reuse. A
more detailed description of each chapter is provided below.
• Chapter 2. Need for and Benefits of Nitrogen and Phosphorus Removal provides background
information on sources of nitrogen and phosphorus in wastewater. It reviews the status of
wastewater treatment in the U.S., the impairment of waterways by excessive nutrients,
government and industry initiatives to reduce nutrient pollution, and the additional benefits and
challenges of nutrient removal.
• Chapter 3. Principles of Phosphorus Removal by Chemical Addition describes the available
forms of metal salts and lime and their reactions with phosphorus. It provides a general
description of solids separation process and the effects of various treatment options on sludge
production and handling.
• Chapter 4. Principles of Biological Nitrogen Removal examines the fundamental microbiology
behind nitrification and denitrification including stoichiometrics and kinetics. It discusses
denitrification kinetics with internal and external carbon sources. Simultaneous nitrification-
denitrification and potential impacts on sludge handling are also discussed.
• Chapter 5. Principles of Biological Phosphorus Removal provides a detailed discussion of the
biological phosphorus removal process including kinetics, substrate requirements,
environmental conditions, design and operational considerations, and impacts on sludge
processing and handling.
• Chapter 6. Overview of Nitrogen and Phosphorus Removal Technologies describes the
technologies available for removing nitrogen, phosphorus, or both from wastewater. Diagrams
are provided for most technologies. It presents information on technology performance
including design and operational factors affecting a plant's ability to achieve low effluent
concentrations.
• Chapter 7. Establishing Design Objectives summarizes this critical step in upgrading or
retrofitting an existing WWTP. It provides guidance on establishing design flow rates,
characterizing flow and contaminants in influent wastewater including detailed sampling
methodologies and data verification steps, and setting goals for process reliability, sustainability,
and flexibility. The chapter also describes solids handling options and site constraints.
• Chapter 8. Selecting Candidate Treatment Processes for Plant Upgrades describes technology
selection factors including treatment goals, available footprint, hydraulic considerations,
chemical needs, solids processing capabilities, and energy considerations. It summarizes
advantages and disadvantages of different technology types. It also provides an overview of a
recommended approach to technology selection and discusses use of advanced tools.
Nutrient Control Design Manual 1-5 August 2010
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• Chapter 9. Design Approach for Phosphorus Removal by Chemical Addition provides guidelines
on selecting a chemical precipitant, choosing application points, and determining chemical dose.
It provides detailed guidance on designing a chemical feed system and considerations for rapid
mix, flocculation, and solids separation processes to maximize phosphorus removal.
• Chapter 10. Design Approach for Biological Nutrient Removal presents a step-by-step approach
for designing wastewater treatment upgrades for nutrient removal using mathematical models.
It provides practical recommendations for data collection and evaluation and model calibration.
It includes design checks for nitrogen and phosphorus removal. This chapter also provides an
alternative design approach using hand or spreadsheet calculations that designers can use to
prepare rough estimates and/or to check model outputs.
• Chapter 11. Design Approach for Effluent Filtration discusses the options in filtration
technology for effluent polishing and nutrient removal. It provides design guidance on granular
media filters and alternative technologies such as cloth filters, disk filters, and membranes.
Information on emerging filtration technologies for removal of phosphorus to low effluent
concentrations is also provided.
• Chapter 12. Operation and Optimization to Enhance Nutrient Removal includes information on
how to optimize the performance of existing operations by incorporating SCADA and other
instrumentation. The chapter also discusses common operational changes to improve system
performance and enhance the cost effectiveness of treatment processes.
• Chapter 13. Instrumentation and Controls discusses online instrumentation for nutrient control
including automated control and optimization, advanced automated control, and SCADA
equipment, all of which can lead to better process optimization and more stable technology
performance.
• Chapter 14. Sustainable Nutrient Recovery and Reuse examines the latest advances in nutrient
recovery including separating and treating waste on-site and how to use wastewater treatment
byproducts to decrease energy costs and greenhouse gas emissions, take advantage of excess
capacity, and generate new revenue.
The manual is supported by three technical appendices containing recommendations on
methanol safety (Appendix A), a list of organic compounds and inhibitory concentrations to nitrification
(Appendix B), and background information on mathematical models for biological nutrient removal
(Appendix C).
1.4 References
Barnard, J. L 1973. Biological Denitrification. Journal of Water Pollution Control 72(6):705-720.
Barnard, J. L. 1975. Nutrient Removal in Biological Systems. Journal of Water Pollution Control, 143-154.
Barnard, J. L. 1978. The Bardenpho Process. In: Advances in Water and Wastewater Treatment Biological
Nutrient Removal. M.P. Wanielista and W.W. Eckenfelder, Jr. (eds.), Ann Arbor Science, Ann Arbor, Ml.
Nutrient Control Design Manual 1-6 August 2010
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Fuhs, G.W. and M. Chen. 1975. Microbiological Basis of Phosphate Removal in the Activated Sludge
Process for the Treatment of Wastewater. Microbial Ecology. 2(2): 119-38
Levin, G.V. and J. Shapiro. 1965. Metabolic Uptake of Phosphorus by Wastewater Organisms. Journal of
the Water Pollution Control Federation (JWPCF), 37 (6): 800-821.
Ludzack, F. J., and M. B. Ettinger. 1961. Controlling Operation to Minimize Activated Sludge Effluent
Nitrogen. Journal of the Water Pollution Control Federation (JWPCF), 34(9):920-931, 1961
Wuhrman, K. 1964. Nitrogen Removal in Sewage Treatment Processes. Verhandlungenden
Internationalen Verein Limnologie, 15:580-596, 1964.
USEPA. 1971. Process Design Manual for Phosphorus Removal. October 1971.
USEPA. 1975. Process Design Manual for Nitrogen Control. October 1975.
USEPA. 1976. Process Design Manual for Phosphorus Removal. EPA/625/l-76-001a, April 1976.
USEPA. 1987a. Design Manual - Phosphorus Removal. EPA/625/1-87/001, September 1987.
USEPA. 1987b. Handbook - Retrofitting POTWsfor Phosphorus Removal in the Chesapeake Bay Drainage
Basin. EPA/625/6-87/017, September 1987.
USEPA. 1993. Manual-Nitrogen Control. EPA/625/R-93/010, September 1993.
USEPAb. 2008a. Municipal Nutrient Removal Technologies Reference Document. Office of Wastewater
Management, Municipal Support Division. EPA 832-R-08-006. Available online:
http://www.epa.gov/OWM/mtb/publications.htm
USEPA. 2008b. Emerging Technologies for Wastewater Treatment and In-Plant Wet Weather
Management. EPA 832-R-06-006. Available online: http://www.epa.gov/OW-
OWM.html/mtb/emerging technologies.pdf
USEPA. 2009. Nutrient Control Design Manual: State of Technology Review Report. Office of Research
and Development. EPA/600/R-09/012. January 2009. Available online at
http://www.epa.gov/nrmrl/pubs/600r09012/600r09012.pdftf22
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2. Need for and Benefits of Nitrogen and Phosphorus Removal
Chapter 2 covers:
2.1 Introduction
2.2 Sources of Nitrogen and Phosphorus in Wastewater
2.3 Status of Wastewater Treatment in the United States
2.4 Nutrient Impairment of U.S. Waterways
2.5 Climate Change Impacts
2.6 Federal and State Regulations and Initiatives to Reduce Nutrient
Pollution
2.7 Industry Initiatives—the WERF Removal Challenge
2.8 Benefits of Nutrient Removal
2.9 Challenges of Nutrient Removal
2.10 References
2.1 Introduction
The harmful effects of eutrophication due to excessive nitrogen and phosphorus concentrations
in the aquatic environment have been well documented. Algae and phytoplankton growth can be
accelerated by higher concentrations of nutrients, leading to harmful algal blooms, hypoxia, and loss of
submerged aquatic vegetation (SAV). Depending on the specific water body characteristics, either
nitrogen or phosphorus can be limiting (i.e., present in the smallest amount compared to growth
requirements). In addition to stimulating eutrophication, nitrogen in the form of ammonia can exert a
direct demand on dissolved oxygen (DO) and can be toxic to aquatic life. Even if a wastewater
treatment plant (WWTP) converts ammonia to nitrate by a biological nitrification process, the resultant
nitrate can stimulate algae and phytoplankton growth.
From a public health perspective, eutrophication may also cause risks to human health, resulting
from consumption of shellfish contaminated with algal toxins or direct exposure to waterborne toxins.
Eutrophication, in particular, can create problems if the water is used as a source of drinking water.
Chemicals used to disinfect drinking water will react with organic compounds in the source water to
form disinfection byproducts, which are potential carcinogens and are regulated by the USEPA. Excess
levels of nitrates above the maximum contaminant level (MCL) in drinking water (10 ppm) can cause
numerous negative health effects due to the body's conversion of nitrate to nitrite, including serious
illness and sometimes death. Infants in particular are susceptible to these effects, which can interfere
with the oxygen-carrying capacity of the blood. This interference can lead to an acute condition in which
health deteriorates rapidly over a period of days. Symptoms include shortness of breath and blueness of
the skin (methemoglobinemia; also known as "Blue-baby Syndrome").
For these reasons, it is important to limit nitrogen and phosphorus contamination of surface and
ground water. One way to minimize this contamination is to reduce levels of nitrogen and phosphorus
in wastewater treatment plant effluent. This chapter will discuss the various sources of these
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contaminants, the impacts they have in the environment, initiatives that are being taken to reduce these
pollutants, and the benefits that can be realized by these efforts.
2.2 Sources of Nitrogen and Phosphorus in Wastewater
This section provides an overview of the sources of nitrogen and phosphorus in wastewater.
2.2.1 Nitrogen
Nitrogen is an essential nutrient for plants and animals. Approximately 80 percent of the Earth's
atmosphere is composed of nitrogen, and it is a key element of proteins and cells. The major
contributors of nitrogen to wastewater are human activities such as food preparation, showering, and
waste excretion. The per capita contribution of nitrogen in domestic wastewater is about one-fifth of
that for biochemical oxygen demand (BOD). Total nitrogen in domestic wastewater typically ranges
from 20 to 70 mg/L for low to high strength wastewater (Tchobanoglous et al. 2003). Factors affecting
concentration include the extent of infiltration and the presence of industries. Influent concentration
varies during the day and can vary significantly during rainfall events, as a result of inflow and infiltration
to the collection system.
The most common forms of nitrogen in wastewater are:
• Ammonia (NH3)
• Ammonium ion (NH4+)
• Nitrite (NO2~)
• Nitrate (NO3~)
• Organic nitrogen
Total Kjeldahl Nitrogen (TKN) is a common nitrogen measurement parameter. This measurement
combines ammonium and organic nitrogen, i.e. reduced forms of nitrogen. It also is typically equal to
the total nitrogen (TN) in wastewaters influent to sewage treatment plants because conditions in sewers
usually result in the reduction of all oxidized forms of nitrogen. However, this is not true in some
collection systems, notably those in steep terrains where sheet flow may dominate flow in the sewers,
resulting in aeration of the sewage and the formation of nitrates during flow. The oxidized forms of
nitrogen must be measured in addition to TKN to determine influent TN for such collection systems.
Nitrogen in domestic wastewater consists of approximately 60 to 70 percent ammonia-nitrogen
and 30 to 40 percent organic nitrogen (Tchobanoglous et al. 2003; Crites and Tchobanoglous 1998).
Most of the ammonia-nitrogen is derived from urea, which breaks down rapidly to ammonia in
wastewater influent.
WWTPs designed for nitrification and denitrification can remove 80 to 95 percent of inorganic
nitrogen, but the removal of organic nitrogen is typically much less efficient (Pehlivanoglu-Mantas and
Sedlak, 2006). Domestic wastewater organic nitrogen may be present in particulate, colloidal, or
dissolved forms and consist of proteins, amino acids, aliphatic N compounds, refractory natural
compounds in drinking water (e.g., humic substances), or synthetic compounds (e.g., ethylene diamine
tetraacetic acid (EDTA) and textile dyes). Organic nitrogen may be released in secondary treatment by
microorganisms either through metabolism or upon death and lysis. Some nitrogen may be contained in
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recondensation products. Hydrolysis of particulate and colloidal material by microorganisms releases
some organic nitrogen as dissolved, biodegradable compounds. Amino acids are readily degraded
during secondary biological treatment, with 90 to 98 percent removal in activated sludge systems and
76 to 96 percent removal in trickling filters. However, other forms of organic nitrogen may be more
persistent in wastewater treatment processes.
The importance of the organic nitrogen fraction has increased as effluent limits on nitrogen
have become more stringent. With more impaired waterways from nutrient loads, effluent limits for
total nitrogen (TN) concentrations of 3.0 mg/L or less are becoming more common. The dissolved
organic nitrogen (DON) concentration in the effluent from biological nutrient removal (BNR) treatment
facilities was found to range from 0.50 to 1.50 mg/L in 80 percent of 188 plants reported by Pagilla
(STAC-WERF 2007), and values as high as 2.5 mg/L were observed. Thus, for systems without effluent
filtration or membrane bioreactors (MBRs) that are trying to meet a TN treatment goal of 3.0 mg/L, the
effluent DON contribution can easily be 20 to 50 percent of the total effluent nitrogen concentration,
compared to only about 10 percent for conventional treatment (Pehlivanoglu-Mantas and Sedlak 2004).
See Chapter 4 of this manual for additional discussion of effluent DON and its implications for nutrient
removal.
2.2.2 Phosphorus
Total phosphorus (TP) in domestic wastewater typically ranges between 4 and 8 mg/L but can be
higher depending on industrial sources, water conservation, or whether a detergent ban is in place.
Sources of phosphorus are varied. Some phosphorus is present in all biological material, as it is an
essential nutrient and part of a cell's energy cycle. Phosphorus is used in fertilizers, detergents, and
cleaning agents and is present in human and animal waste.
Phosphorus in wastewater is in one of three forms:
• Phosphate (also called Orthophosphate)
• Polyphosphate
• Organic phosphorus
The orthophosphate fraction is soluble and can be in one of several forms (e.g., phosphoric acid,
phosphate ion) depending on the solution pH. Polyphosphates are high-energy, condensed phosphates
such as pyrophosphate and trimetaphosphate. They are also soluble but will not be precipitated out of
wastewater by metal salts or lime. They can be converted to phosphate through hydrolysis—which is
very slow—or by biological activity.
Organic phosphorus can be soluble, colloidal or particulate, i.e. settleable. It can also be divided
into biodegradable and non-biodegradable fractions. Particulate and colloidal organic phosphorus is
generally settled or precipitated out and removed with the sludge. Soluble organic biodegradable
phosphorus can be hydrolyzed into orthophosphate during the treatment process. Soluble organic non-
biodegradable phosphorus will pass through a WWTP. Assuming an influent TP of 6 -8 mg/L, a typical
wastewater will contain 3 to 4 mg/L phosphorus as phosphate, 2 to 3 mg/L as polyphosphate, and 1
mg/L as organically bound phosphorus (WEF and ASCE 2006).
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2.3 Status of Wastewater Treatment in the United States
The 1972 Amendments to the Federal Water Pollution Control Act (FWPCA) (Public Law 92-500),
also known as the Clean Water Act (CWA), established the foundation for wastewater discharge control
in the United States. The CWA's primary objective is to "restore and maintain the chemical, physical,
and biological integrity of the Nation's waters." The CWA established a program to ensure clean water
by requiring permits that limit the amount of pollutants discharged by all municipal and industrial
dischargers into receiving waters. Discharges are regulated under the National Pollutant Discharge
Elimination System (NPDES) permit program. As of 2004, there were 16,583 municipal wastewater
treatment plants [also known as Publicly Owned Treatment Works (POTWs)] regulated under the CWA,
serving approximately 75 percent of the nation's population (U.S. Public Health Service and USEPA 2008)
with the remaining population served by septic or other decentralized (cluster) systems.
Wastewater treatment has generally been defined as containing one or more of the following
four processes: (1) preliminary, (2) primary, (3) secondary, and (4) advanced (or tertiary) treatment.
Preliminary treatment consists of grit removal, which removes dense inert particles, and screening to
remove rags and other large debris. Primary treatment involves gravity settling tanks to remove
settleable solids, including settleable organic solids. The performance of primary settling tanks can be
enhanced by adding chemicals to capture and flocculate smaller solid particles for the precipitation and
removal of phosphorus. Secondary treatment follows primary treatment in most plants and employs
biological processes to remove colloidal and soluble organic matter. Effluent disinfection is usually
included in the definition of secondary treatment.
EPA classifies advanced treatment as "a level of treatment that is more stringent than secondary
or produces a significant reduction in conventional, non-conventional, or toxic pollutants present in the
wastewater" (U.S. Public Health Service and USEPA 2008). Other technical references subdivide
advanced treatment, using the terms "secondary with nutrient removal" when nitrogen, phosphorus, or
both are removed and "tertiary removal" to refer to additional reduction in solids by filters or
microfilters (Tchobanoglous et al. 2003). Effluent filtration and nutrient removal are the most common
advanced treatment processes.
The CWA requires that all municipal WWTP discharges meet a minimum of secondary
treatment. Based on data from the 2004 Clean Watersheds Needs Survey, 16,543 municipal WWTPs
(99.8 percent of plants in the country) meet the minimum secondary wastewater treatment
requirements. Of those that provide at least secondary treatment, approximately 44 percent (7,322
plants) provide some kind of advanced treatment (U.S. Public Health Service and USEPA 2008). Figure 2-
1 shows how secondary and advanced wastewater treatment has been implemented since 1940 and
also provides projected treatment for 2024. Note that "No Discharge" refers to systems that do not
discharge treated wastewater to the nation's waterways but instead dispose of wastewater via methods
such as industrial reuse, irrigation, or evaporation.
Nutrient Control Design Manual 2-4 August 2010
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100.0% -r
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"5 ^5
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• Secondary
• Less Than Secondary
• Raw
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1950 1962 1968 1972 1978 1982 1988 1992 1996 2000 2004 Projected
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Figure 2-1. Population served by POTWs nationwide for select years between 1940 and 2004 and projected to 2024 (if all needs
are met), organized by wastewater treatment type. Source: U.S. Public Health Service and USEPA Clean Watersheds Needs
Surveys 2004 Report to Congress (U.S. Public Health Service and USEPA 2008).
2.4 Nutrient Impairment of U.S. Waterways
According to the 2007 report Effects of Nutrient Enrichment in the Nation's Estuaries: A Decade
of Change, increased nutrient loadings promote a progression of symptoms that begin with excessive
growth of phytoplankton and macroalgae and advance to the point where grazers cannot control
growth (Bricker et al., 2007). These blooms may be problematic, potentially lasting for months at a time
and blocking sunlight to light-dependent SAV. In addition to increased growth, changes in naturally
occurring ratios of nutrients may also affect which species dominate, potentially leading to
nuisance/toxic algal blooms. These blooms may also lead to other, more serious symptoms that affect
biota, such as low DO and loss of SAV. Once water column nutrients have been depleted by
phytoplankton and macroalgae and these blooms die, the bacteria decomposing the algae then
consume oxygen, making it less available to surrounding aerobic aquatic life. Consequently, fish and
invertebrate kills may occur due to hypoxia and anoxia (conditions of low to no DO).
Four examples of impaired large water bodies impacted by nutrient loadings are described
below. There are more than 80 additional estuaries and bays, and thousands of rivers, streams, and
lakes that are also impacted by nutrients in the United States. In fact, all but one state and two
territories have CWA section 303(d) listed1 water body impairments for nutrient pollution. Collectively,
states have listed over 10,000 nutrient and nutrient-related impairments.
2.4.1 Northern Gulf of Mexico
Advanced eutrophic conditions can lead to "dead zones" with limited aquatic life, which
describes the hypoxia condition that exists in the Northern Gulf of Mexico. A recent U.S. Geological
Survey (USGS) report titled Differences in Phosphorus and Nitrogen Delivery to the Gulf of Mexico from
1 Required by Section 303(d) of the CWA, the 303(d) list is a list of state's water bodies that do not meet or are not
expected to meet applicable Water Quality Standards with technology-based controls alone.
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the Mississippi River Basin documents the contribution of nitrogen and phosphorus from agricultural
and non-agricultural sources in the Mississippi River basin (Alexander et al. 2008). On June 16, 2008, the
joint federal-state Mississippi River/Gulf of Mexico Watershed Nutrient Task Force released its 2008
Action Plan for Reducing, Mitigating, and Controlling Hypoxia in the Northern Gulf of Mexico and
Improving Water Quality in the Mississippi River Basin, which builds upon its 2001 plan by incorporating
emerging issues, innovative approaches, and the latest science, including findings from EPA's Science
Advisory Board. Improvements include more accountability through an Annual Operating Plan, better
tracking of progress, state and federal nutrient reduction strategies, and a plan to increase awareness of
the problem and implementation of solutions (USEPA 2008c).
2.4.2 Chesapeake Bay
Nutrient pollution has caused significant problems in the Chesapeake Bay. Elevated levels of
both nitrogen and phosphorus are the main causes of poor water quality and loss of aquatic habitats in
the Bay. Significant algae blooms on the water surface block the sun's rays from reaching underwater
bay grasses. Without sunlight, bay grasses cannot grow and provide critical food and habitat for blue
crabs, waterfowl, and juvenile fish. Created in 1983, the Chesapeake Bay Program (CBP) is a
comprehensive cooperative effort by federal, state, and local governments; nongovernmental
organizations; academics; and other entities that share the mission of restoring and protecting the
Chesapeake Bay and its watershed. The CBP estimates that 22 percent of the phosphorus loading and 19
percent of the nitrogen loading in the Bay comes from municipal and industrial wastewater facilities
(Chesapeake Bay Program, 2008).
To address the need to decrease these pollutants and restore the health of the Bay, the US EPA
Chesapeake Bay Office (CBO), along with its partners, developed the Chesapeake Action Plan (CAP),
which included goals for the year 2000. The CAP provided a strategic framework that unified CBP's
existing planning documents and clarified how CBP partners would pursue the restoration and
protection goals for the Bay and its watershed; created an activity integration plan with comprehensive,
quality assured data for 2007 that identifies and catalogs CBP partners' implementation activities and
corresponding resources; and provided high-level summaries of key information, such as clear status of
progress, expected progress toward certain Chesapeake 2000 goals, summaries of actions and funding,
and a brief summary of the challenges and actions needed to expedite progress. A summary of the
progress to date can be found in a July 2008 Report to Congress: Strengthening the Management,
Coordination, and Accountability of the Chesapeake Bay Program. To date, uniform water quality
standards have been adopted across the Bay, an aggressive permitting approach for WWTPs has been
established, the most cost-effective agricultural best management practices (BMPs) have been
implemented, and forests and wetlands surrounding the Bay have been restored.
2.4.3 Great Lakes
The Great Lakes were the first bodies of water to attract national attention to the problem of
nutrient caused eutrophication. In the 1960s, Lake Erie was declared "dead" when excessive nutrients
in the Lake stimulated excessive algae blooms that covered beaches and killed off native aquatic species
as a result of oxygen depletion. At that time, phosphorus was the primary nutrient of concern due to
the use of phosphate detergents and inorganic fertilizers. Algal assays showed that phosphorus was the
limiting nutrient in the Lakes, which is typical for fresh water bodies. Thus, the focus was on the control
of phosphorus. With the enactment of the CWA and the Great Lakes Water Quality Agreement in 1972,
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a concerted effort was undertaken to reduce pollutant loadings, particularly phosphorus, to the Lake.
Although the health of the Lake improved dramatically, in recent years, there has been renewed
attention to the re-emergence of a "dead" zone in Lake Erie, again due to nutrient loadings. Recent
studies by scientists and the National Oceanic and Atmospheric Administration (NOAA) have also
hypothesized a relationship between excessive nutrients in the Lake and the presence of two aquatic
invasive species—the zebra mussel and the quagga mussel (Vanderploeg et al. 2002).
2.4.4 Long Island Sound
Development and population increases in the Long Island Sound Watershed have resulted in a
significant increase in nitrogen loading, the current limiting nutrient to the Sound. The increased
nitrogen loads have stimulated plant growth, increased the amount of organic matter settling to the
benthic zone, lowered DO levels, and changed habitats. The primary concerns in the Sound include
hypoxia, the loss of sea grass, and alterations in the food web. Management efforts are currently
underway to reduce nitrogen pollution by more than half with a focus on two areas: (1) upgrading
WWTPs with new technologies and (2) removing nitrogen by reducing polluted run-off through BMPs on
farms and suburban areas (Long Island Sound Study 2004). The emphasis has been primarily on reducing
nitrogen discharges, but in recent years phosphorus limitations have been placed on upstream WWTPs
discharging to rivers that discharge into the Sound.
2.5 Climate Change Impacts
Climate change may also be a significant influence on the development of future eutrophic
symptoms. According to the report Effects of Nutrient Enrichment in the Nation's Estuaries: A Decade of
Change, the factors associated with climate change that are expected to have the greatest impacts on
coastal eutrophication are:
• Increased temperatures
• Sea level rise
• Changes in precipitation and freshwater runoff
Increased temperatures will have several effects on coastal eutrophication. Most coastal species
are adapted to a specific range of temperatures. Increases in water temperatures may lead to expanded
ranges of undesirable species. Higher temperatures may also lead to increased algal growth and longer
growing seasons, potentially increasing problems associated with excessive algal growth and
nuisance/toxic blooms. Additionally, warmer waters hold less DO, therefore potentially exacerbating
hypoxia. Temperature-related stratification of the water column may also worsen, having a further
negative effect on DO levels.
Climate change models predict increased melting of polar icecaps and changes in precipitation
patterns, leading to sea level rise and changes in water balance and circulation patterns in coastal
systems. Sea level rise will gradually inundate coastal lands, causing increased erosion and sediment
delivery to water bodies, and potentially flooding wetlands. The increased sediment load and
subsequent turbidity increase may cause SAV loss. As erosion increases, sediment-associated nutrients
also increase, stimulating algal growth. This positive feedback between increased erosion and algal
growth may also increase turbidity. The loss of wetlands, which act as nutrient sinks, will further
increase nutrient delivery to estuaries. In contrast, an increase in freshwater inflow may reduce
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residence time of pollutants, reducing the probability of blooms in some systems. In regions of
engineered water flow (e.g., South Atlantic and Gulf of Mexico), the impacts of changes in the amount of
runoff will depend on how water management strategies control regional hydrology.
Another report titled Aquatic Ecosystems and Global Climate Change — Potential Impacts on
Inland Freshwater and Coastal Wetland Ecosystems in the United States notes that climate change of the
magnitude projected for the United States over the next 100 years will cause significant changes to
temperature regimes and precipitation patterns across the nation (Poff et al. 2002). Such alterations in
climate pose serious risks for inland freshwater ecosystems (lakes, streams, rivers, wetlands) and coastal
wetlands, and may adversely affect numerous critical services provided to human populations.
2.6 Federal and State Regulations and Initiatives to Reduce Nutrient Pollution
This section describes the following Federal and state regulations and initiatives to reduce
nutrient pollution:
• Water Quality Standards (WQSs)
• Total Maximum Daily Loads (TMDLs)
• National Pollution Discharge Elimination System (NPDES) Permitting
• Water quality trading
• Technology evaluations and guidance
In addition to these efforts, state and EPA surface water and drinking water program managers formed
an ad hoc Nutrient Innovations Task Group in October 2008 to "identify and frame key nutrient issues,
questions, and options on how to improve and accelerate nutrient pollution prevention and reduction
at the state and national level" (State-EPA Nutrient Innovations Task Group, 2009). Their final report is
available online at http://www.epa.gov/waterscience/criteria/nutrient/ .
2.6.1 Water Quality Standards
Water quality standards (WQS) are the foundation of the water quality-based pollution control
program mandated by the CWA. WQS define the goals for a water body by designating its uses, setting
criteria to protect those uses, and establishing provisions to protect water bodies from pollutants. The
WQS regulation requires that states and authorized Indian tribes specify appropriate water uses to be
achieved and protected. Appropriate uses are identified by considering the use and value of the water
body for public water supply; for protection of fish, shellfish, and wildlife; and for recreational,
agricultural, industrial, and navigational purposes. In designating uses for a water body, states and tribes
examine the suitability of a water body for the uses based on the physical, chemical, and biological
characteristics of the water body; its geographical setting and scenic qualities; and economic
considerations.
States and tribes typically adopt both numeric and narrative criteria. Numeric criteria are
important where the cause of toxicity is known or for protection against pollutants with potential
human health effects. Narrative criteria are also important—narrative "free from" toxicity criteria
typically serve as the basis for limiting the toxicity of waste discharges to aquatic species. In addition to
Nutrient Control Design Manual 2-8 August 2010
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narrative and numeric (chemical-specific) criteria, states can and often adopt nutrient criteria, which are
defined as a means to protect against nutrient over-enrichment and eutrophication.
In 1998, EPA published a National Strategy for the Development of Regional Nutrient Criteria, a
roadmap for designing guidance for states to use in the development of numeric water quality criteria
for nutrients. These technical guidance documents describe the techniques used to develop nutrient
criteria for use in state and tribal water quality standards. They cover:
• Estuarine and Coastal Waters
• Lakes and Reservoirs
• Rivers and Streams
• Wetlands
In addition to technical guidance for developing nutrient criteria, EPA has published additional
technical documents and tools to aid states and tribes in assessing nutrients in their waters. These
include:
• Methods for Evaluating Wetland Conditions: A series of documents to help states and tribes
build their capacity to monitor and assess the biological and nutrient conditions of wetlands.
These modules are designed to help states and tribes establish biological and nutrient
assessment and monitoring programs for wetlands. These documents are available online at
http://www.epa.gov/waterscience/criteria/nutrient/guidance/wetlands/tfmodules
• N-STEPS: A database that provides technical assistance to state and regional scientists and
managers who are developing numeric nutrient criteria and provides information to the public
regarding nutrient pollution and EPA's activities. This database is available online at
http://www.epa.gov/waterscience/criteria/nutrient/n-steps.html
• Other Nutrient Databases: Databases developed by EPA that contain sample data from various
waters throughout the United States, available online at
http://www.epa.gov/waterscience/criteria/nutrient/database/index.html
Numeric nutrient water quality standards drive water quality assessments and watershed
protection management. They support improved development of nutrient Total Maximum Daily Loads
(TMDLs; described below). Perhaps most importantly, they create state- and community-developed
environmental baselines that allow states and EPA to manage more effectively, measure progress, and
support broader partnerships based on nutrient trading, BMPs, land stewardship, wetlands protection,
voluntary collaboration, and urban storm water runoff control strategies.
There are a number of key advantages to adopting numeric standards, including:
• Easier and faster development of TMDLs
• Quantitative targets to support trading programs
• Easier to write protective NPDES permits
• Increased effectiveness in evaluating success of nutrient runoff minimization programs
• Measurable, objective water quality baselines against which to measure environmental progress
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In December 2008, EPA published a report titled, State Adoption of Numeric Nutrient Standards
(1998-2008). The report, which can be found at
http://www.epa.Rov/waterscience/criteria/nutrient/files/reportl998-2008.pdf, provides an update of
state efforts to develop numeric nutrient standards.
Figure 2-2 depicts the relationships among WQS and the tools used to help meet these
standards such as TMDLs, NPDES permits, and water quality trading. These tools are discussed in more
detail below.
Defines the water
quality goal
Compile data/information
and assess vvaterbody
condition
303(d) Program
40 CFR 130.7
Implementation
Manage Naffpoini Sources
Thrciixjh L'dranln. PalnnrsNpK
and Voluntary Prugranis
Figure 2-2. Water quality-based approach of the Clean Water Act
Source: USEPA2008f.
2.6.2 Total Maximum Daily Loads (TMDLs)
CWA Section 303(d) requires states to develop TMDLs for water bodies on the 303(d) list of
impaired waters (e.g., waters not meeting their designated uses). A TMDL is a calculation of the
maximum amount of a pollutant a water body can receive and still meet WQS. TMDLs serve as a tool for
implementing WQS. The TMDL targets or endpoints represent a number where the applicable WQS and
designated uses (e.g., public water supply, contact recreation, and the propagation and growth of
aquatic life) are achieved and maintained in the water body of concern. TMDLs identify the level of
pollutant control necessary to meet WQS and support the designated uses of a water body. Once a
TMDL is set, the total load is allocated among all existing sources. The allocation is divided into two
portions: (1) a load allocation (LA) representing natural and non-point sources and (2) a waste load
allocation (WLA) representing NPDES-permitted point source discharges. In many regions, water bodies
have a poor ability to assimilate nutrients or they are already impaired from past pollution and cannot
handle large loads of additional nutrients. In these cases, TMDLs may require nutrient permit levels to
be even lower than what might otherwise be allowed by nutrient criteria.
Although states are not required under section 303(d) to develop TMDL implementation plans,
many states include implementation plans with the TMDL or develop them as a separate document.
When developed, TMDL implementation plans may provide additional information on what point and
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nonpoint sources contribute to the impairment and how those sources are being controlled, or should
be controlled in the future. Once a TMDL has been defined and various load allocations established, the
CWA requires that all NPDES permits associated with the water body must reflect the loads established
in the TMDL. For WWTPs, this may include specific criteria for nutrients.
2.6.3 NPDES Permitting
Established by the CWA Amendment of 1972, EPA's NPDES permit program has been the
primary mechanism for controlling pollution from point sources. Point sources are discrete conveyances
such as pipes or man-made ditches. Individual homes that are connected to a municipal system, use a
septic system, or do not have a surface discharge do not need an NPDES permit; however, POTWs and
other facilities must obtain permits if they discharge directly to surface waters.
NPDES permits for wastewater discharges contain, among other information, effluent limits for
"conventional" pollutants such as BOD, total suspended solids (TSS), and pH as well as limits for specific
toxicants including various organic and inorganic chemicals. Permits may also include effluent limits for
"non-conventional" pollutants such as nitrogen and phosphorus. Effluent limits can be technology-
based and/or water-quality based. EPA has established technology-based, secondary treatment effluent
limits for BOD as 5-day biochemical oxygen demand (BOD5), TSS, and pH. Water-quality based effluent
limits are set if the technology-based limits are not sufficient to maintain the WQS of the receiving
water.
In December 2003, EPA published the Watershed-Based National Pollutant Discharge
Elimination System (NPDES) Permitting Implementation Guidance
(http://www.epa.gov/npdes/pubs/watershedpermitting finalguidance.pdf), which describes the
concept of and the process for watershed-based permitting under the NPDES permit program.
Watershed-based NPDES permitting is an approach to developing NPDES permits for multiple point
sources located within a defined geographic area (watershed boundaries) to meet WQS. This approach,
aimed at achieving new efficiencies and environmental results, provides a process for considering all
stressors within a hydrologically defined drainage basin or other geographic area, rather than addressing
individual pollutant sources on a discharge-by-discharge basis. This report was followed by technical
guidance issued in August 2007 titled, Watershed-based National Pollutant Discharge Elimination
System (NPDES) Permitting Technical Guidance, available online at
http://www.epa.gov/npdes/pubs/watershed techguidance entire.pdf
Federal and State regulations related to WQS and TMDLs as described previously are expected
to result in more stringent NPDES effluent limits for nitrogen and phosphorus.
2.6.4 Water Quality Trading
Water quality trading is a market-based approach to improve and preserve water quality once
WQS and/or TMDLs have been defined. Trading can provide greater efficiency in achieving water quality
goals by allowing one source to meet its regulatory obligations by using pollutant reductions created by
another source that has lower pollution control costs. For example, under a water quality trading
program, a POTW could comply with discharge requirements by paying distributed sources to reduce
their discharges by a certain amount. The use of geographically-based trading ratios provides an
economic incentive, encouraging action toward the most cost-effective and environmentally beneficial
projects.
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EPA issued a Water Quality Trading Policy in 2003 to provide guidance to states and tribes on
how trading can occur under the CWA and its implementing regulations
(http://www.epa.gov/owow/watershed/trading/finalpolicy2003.pdf). The policy discusses CWA
requirements that are relevant to water quality trading including: requirements to obtain permits; anti-
backsliding provisions; development of WQS, including an anti-degradation policy; NPDES permit
regulations, TMDLs; and water quality management plans. EPA also developed a number of tools and
guidance documents to assist states, permitted facilities, non-point sources, and stakeholders involved
in the development of trading programs (www.epa.gov/owow/watershed/trading.htm). Recently, the
U.S. Department of Agriculture (USDA) National Resources Conservation Service released a Nitrogen
Trading Tool (NTT) prototype for calculating nitrogen credits based on the Nitrogen Loss and
Environmental Assessment Package Model (Gross et al. 2008).
Water quality trading programs have been successfully implemented in several states and
individual watersheds across the county. For example, nitrogen pollution from point sources into the
Long Island Sound was reduced by nearly 25 percent using an innovative nitrogen credit trading
program. In Connecticut, the program was implemented across 79 sewage treatment plants in the
state. Through the Nitrogen Credit Exchange, established in 2002, the Connecticut program has a goal of
reducing nitrogen discharges by 58.5 percent by 2014.
A recent American Society of Civil Engineers journal article (Landers 2008) points out, however,
that regulatory frameworks for water quality trading programs have yet to be adopted by the majority
of states. Barriers to adopting such programs include uncertainty in: (1) the mechanisms for
determining appropriate credits and ratios between point sources and distributed sources; and (2)
approaches to ensure that promised reductions actually occur. Other barriers include lack of
resources/staff to organize such programs and lack of specific nutrient goals/TMDLs to drive markets.
2.6.5 Technology Evaluation and Guidance
In addition to regulatory and policy initiatives, EPA helps control nutrients through the
development and dissemination of technical information. For example, EPA's Office of Wastewater
Management (OWM) has developed a number of technology fact sheets on secondary and advanced
biological treatment (USEPA, 1999b; 1999c). OWM has also published several technology reports,
including Emerging Technologies for Wastewater Treatment and In-Plant Wet Weather Management
(USEPA, 2008a). This technology guide, published in February 2008, is designed to help municipal
wastewater treatment system owners and operators find information on emerging wastewater
treatment and in-plant wet weather management. In September 2008, OWM published the Municipal
Nutrient Removal Technologies Reference Document (USEPA 2008d). Volume 1 of this report provides
detailed technical and cost information on biological and physiochemical treatment technologies for the
removal of nitrogen, phosphorus, or a combination of the two. The report also includes at least one
year's worth of full-scale performance data for 27 wastewater treatment facilities in the United States
and Canada and 9 detailed case studies. EPA's Region 10 initiated a project to evaluate municipal
WWTPs that have demonstrated exemplary phosphorus removal through their treatment processes. In
April 2007, the Region published a report titled, Advanced Wastewater Treatment to Achieve Low
Concentration of Phosphorus (EPA Region 10 2007).
In 1975, EPA's Office of Research and Development (ORD) published its first technology design
guidance for nitrogen removal: Process Design Manual for Nitrogen Control. The manual was updated in
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1993 and focused on biological/mechanical processes that were finding widespread application for
nitrification and nitrogen removal at that time. The development of guidance for phosphorus removal
followed a similar schedule, with ORD publishing the document Process Design Manual for Phosphorus
Removal in 1971. In 1976, the manual was updated to include design guidance for phosphorus removal
using mineral addition and lime addition. In 1987, EPA published two technical documents to address
phosphorus control: (1) an update to the 1976 Process Design Manual for Phosphorus Removal and (2) a
handbook titled, Handbook — Retrofitting POTWsfor Phosphorus Removal in the Chesapeake Bay
Drainage Basin. EPA has revised these materials to provide updated state-of-the technology design
guidance for both nitrogen and phosphorus control at municipal WWTPs as presented in this manual.
2.7 Industry Initiatives—The WERF Nutrient Removal Challenge
In early 2007, the Water Environment Research Foundation (WERF) created the Nutrient
Removal Challenge program with the goals of:
• Identifying, assessing, and making recommendations to improve sustainable wastewater
nutrient removal technologies.
• Providing information to help agencies meet various receiving water body requirements and
other wastewater treatment goals (e.g., climate change, sustainability, cost-effectiveness,
reliability).
• Conducting research to inform regulatory decision making and help practitioners comply
with increasingly high levels of nitrogen and phosphorus removal with a focus on improving
plant performance.
This multi-year program will be funded for 5 years with WERF and external funds anticipated to total $8-
10 million.
As part of the program's kick-off activities, the WERF Nutrient Research Stakeholder Workshop
was held on March 7 and 8, 2007 in Baltimore, MD, to further refine the Challenge's research needs and
to seek funding partners and collaborators. A total of 25 priority areas were identified, many of them
analogous to those identified in a similar workshop conducted by WERF in 2006. Generally, these
research areas fall into one of the following three categories:
• Characterization of effluent organic nitrogen
• Accuracy of analytical measurement techniques for low concentrations of phosphorus
• Alternative carbon sources for denitrification
WERF will also be developing a Nutrient Compendium, a comprehensive, living document that
will describe the current knowledge of regulatory and technological nutrient removal issues. The
document will detail the key knowledge areas affecting nutrient removal to very low limits and identify
knowledge gaps related to nutrient removal.
As of 2009, WERF has:
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• Launched a new web portal (http://www.werf.org/nutrients ) with research information
and relevant links
• Developed technical documents for the Nutrient Compendium (preparation of additional
technical summaries is underway), including:
- Tertiary phosphorus removal
Low phosphorus concentration measurement
Effluent dissolved organic nitrogen
External carbon augmentation and denitrification
• Initiated collaborative research in three key areas
Effluent organic nitrogen/ refractory dissolved organic nitrogen (RDON)
Low phosphorus analysis and measurement at very low limits
- Alternative external carbon for denitrification to reduce cost and improve efficiency of
nitrogen removal
2.8 Benefits of Nutrient Removal
2.8.1 Improved Plant Performance
Biological nutrient removal at POTWs can lead to several operational improvements. In most
cases, the addition of an anaerobic zone for biological phosphorus removal will increase the sludge
density because of phosphorus accumulation, and reduce the growth of filamentous organisms because
of the absence of DO, thereby improving settleability. A pre-anoxic zone for denitrification can also
lead to a more stable, better settling activated sludge process as the anoxic-aerobic processes favor
good settling floe-forming bacteria over filamentous growth. Additional benefits of a pre-anoxic zone
include less aeration energy required in the aerobic zone as the nitrate produced can be used for BOD
removal, and less sludge production compared to post-anoxic treatment with supplemental carbon.
Anaerobic and anoxic zones also provide better control of foaming if backmixing is eliminated and the
recycle of NOx and DO to the zones is minimized. Further, good removal of nitrogen reduces concern
over denitrification and floating sludge in the secondary clarifiers, and provides the option of oversizing
the clarifiers to better handle wet weather flows.
The addition of an anaerobic or anoxic zone ahead of the aerobic zone can improve the rate of
oxygen transfer in the water. As an example, WEF and ASCE (2006) reported an almost doubling of the
oxygen transfer efficiency when the 23rd Avenue Plant in Phoenix, AZ, added a pre-anoxic zone.
Nitrification in the aerobic zone can significantly reduce the alkalinity of the mixed liquor and, in
situations where the influent alkalinity is low (e.g., a municipality with soft water), can completely
deplete the alkalinity and cause a rapid drop in pH. The nitrification rate will slow significantly below a
pH of approximately 6.8. Tchobanolgous et al. (2003) report that nitrification rates at a pH of 6.0 may
only be about 20 percent of that with a pH of 7.0. Thus, plants with low alkalinity may experience
periods of reduced nitrification and elevated ammonia levels in the plant effluent. Denitrification for
total nitrogen removal can replenish much of the alkalinity lost during nitrification and improve
operational stability. Approximately 62.5 percent of the amount consumed during nitrate formation is
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recovered from the nitrates reduced to nitrogen gas. The total recovery is seldom more than 50 percent
for recycle configurations such as MLE with complete denitrification in the anoxic zone.
2.8.2 Co-Removal of Emerging Contaminants
The term "emerging contaminants" refers broadly to those synthetic or naturally occurring
chemicals, or to any microbiological organisms, that have not been commonly monitored in the
environment but are of increasing concern because of their known or suspected adverse ecological or
human health effects. Several studies have examined the effectiveness of current wastewater
treatment technologies in the removal of emerging contaminants. Significant findings are:
• Removal efficiencies were enhanced for several investigated contaminants at longer solids
residence times (SRTs), with critical SRTs for some contaminants beyond which removal rates
did not improve.
• Longer SRTs allow for the establishment of slower growing bacteria (e.g., nitrifying bacteria in
activated sludge), which in turn provide a more diverse community of microorganisms with
broader physiological capabilities, and the potential for greater biodegradation of emerging
contaminants.
• Reverse osmosis has been found to effectively remove pharmaceutical and personal care
products (PPCPs) below detection limits including those that were not consistently removed at
longer SRTs.
A more detailed discussion of removal of emerging contaminants can be found in EPA's Nutrient
Control Design Manual —State of Technology Report (EPA/600/R-09/012, January 2009) at
http://www.epa.gov/nrmrl/pubs/600r09012/600r09012.pdf.
2.8.3 Nutrient Recovery and Reuse
Nutrient recovery and reuse is gaining national and international attention as a key aspect in
sludge management plans. Rather than being disposed as a "waste," sludge is now being harvested for
valuable resources and used as an alternative source of energy. In February 2008, the Global Water
Research Coalition in cooperation with USEPA and WERF released a report titled State of the Science
Report: Energy and Resource Recovery from Sludge (Kalogo and Monteith 2008). In the report, energy
recovery technologies are classified into sludge-to-biogas, sludge-to-syngas, sludge-to-oil, and sludge-to-
liquid processes. The report also describes various technologies available for resource recovery,
including those to recover phosphorus, building materials, nitrogen, and volatile acids. Refer to Chapter
14 of this Manual for more information on resource recovery and sustainability.
2.9 Challenges of Nutrient Removal
Two potentially negative environmental impacts of employing advanced technologies to remove
nutrients from wastewater are the increase in the energy use and release of nitrous oxide (N2O), a
greenhouse gas, into the atmosphere.
2.9.1 Energy Requirements
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While some operational strategies for enhancing nutrient removal, such as cyclic aeration for
denitrification and the utilization of fixed media for integrated fixed-film activated sludge (IFAS)
operation in partially or fully nitrifying systems, do not require additional electricity to operate, many
BNR technologies require an input of energy beyond that needed for conventional municipal treatment.
This is obviously true if the system to be retrofitted was not designed to nitrify due to additional
aeration required. Other examples of retrofits that require additional energy include (Kang et al. 2009):
Additional mixing basins
Chemical addition
Additional pumping for recycle flows
Additional pumping for filtration
Use of an external carbon source
It is not unusual for the additional energy requirements of upgrading to BNR to require an upgrade to
the entire power system, including physical space for motor control centers.
Nitrification requires 50 to 100 percent more aeration energy than non-nitrifying systems,
depending upon the operating SRT selected. The total BOD-NOD aeration costs, however, will be
reduced by about 20 percent by denitrification to 8-10 mg/LTN for typical municipal wastewaters. The
quantity of WAS produced will be greatly reduced which can substantially reduce the amount of energy
needed for dewatering and disposal. Thus, conversion of a non-nitrifying activated sludge system to
EBPR may not require much in the way of additional energy. Note, also, that addition of either an
anaerobic or anoxic zone (and particularly both) ahead of the aerobic zone reduces the aeration volume
needed for nitrification through removal from solution or through stabilization of BOD ahead of the
aerobic zone, which increases the effective nitrification rate in the aerobic zone.
Chapter 8 of this design manual provides additional discussion of energy and identifies
strategies that can be implemented by nutrient removal plants to reduce energy requirements and
improve sustainability. The remainder of this section identifies EPA tools and other general strategies
that can be used to reduce energy needs.
EPA encourages utilities to identify approaches to integrate energy-efficient practices into their
daily management and long-term planning. The 2008 publication, Ensuring a Sustainable Future, an
Energy Management Guidebook for Wastewater and Water Utilities (USEPA 2008i), provides a
recommended approach for energy management using the Plan-Do-Check-Act approach. This
guidebook is available online at
http://www.epa.gov/waterinfrastructure/bettermanagement energy.html.
A September 2008 publication titled National Water Program Strategy— Response to Climate
Change (USEPA 20008b) outlines a number of steps that EPA has taken and that utilities can take to
improve energy efficiency. For example, EPA's ENERGY STAR program has developed a "Focus" in the
water and wastewater industries. An ENERGY STAR Focus is a targeted effort to improve the energy
efficiency within a specific industry or combination of industries. It creates momentum for continuous
improvement in energy performance, provides the industry's managers with the tools they need to
achieve greater success in their energy management programs, and creates a supportive environment
where energy efficiency ideas and opportunities are shared.
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Significant progress is being made in the development of new tools for benchmarking energy
performance among public water and wastewater utilities. For example, the ENERGY STAR program is
expanding the capability of its Energy Performance Rating System to enable drinking water and
wastewater utilities to assess their energy use over time and compare it to other utilities—normalized
for weather and facility characteristics. As of October 2007, WWTP energy performance can be rated
using the ENERGY STAR program's on-line tool, Portfolio Manager. Portfolio Manager can be used to
establish baseline energy use, prioritize investments, set goals, and track energy use and carbon
emissions reductions over time.
Many facilities have already installed alternative energy power production facilities, including
solar, wind, and hydro, for heating and electricity generation. For example, Calera Creek Water
Recycling Plant in Pacifica, CA is using solar panels that provide 10-15 percent of its energy needs,
resulting in an estimated $100,000 savings annually in energy costs (USEPA 2006).
Wastewater facilities can also generate energy from the capture and use of methane. Combined
heat and power (CHP) systems can recover biogas (a mixture of methane, carbon dioxide, water vapor,
and other gases) from anaerobic digesters to heat buildings or to generate electricity. For example, the
East Bay Municipal Utility District (EBMUD) captures and uses biogas to generate enough energy to
cover 90 percent of energy needed at its main wastewater facility. If all 544 large sewage treatment
plants in the United States operating anaerobic digesters were to install combined heat and power,
about 340 megawatts of clean energy could be generated, offsetting 2.3 million metric tons of carbon
dioxide emissions annually (i.e., equivalent to planting about 640,000 acres of forest, or the emissions of
about 430,000 cars) (USEPA 2007a). This energy is also marketable as "green power" to power utilities
that are now required by State laws to have alternative or "green" power as a part of their overall
production.
2.9.2 Release of Nitrous Oxide
Nitrous oxide (N2O) is considered to be a greenhouse gas with a global warming potential
approximately 300 times that of the contribution of carbon dioxide (CO2). N2O as well as nitric oxide
(NO) are known to be intermediates of heterotrophic denitrification (Ahn et al. 2009). Until recently,
there has been considerable uncertainty regarding the mechanisms of N2O generation and whether it is
produced and released during the nitrification and denitrification processes.
In 2008, WERF initiated a research project (no. U4RO7) to quantify N2O emissions, determine
the mechanisms by which it forms, and develop operational strategies to prevent its formation and
release. The first step of this project was to develop a detailed protocol for measuring N2O and NO
emissions from an activated sludge reactor. This method has now been reviewed by EPA and
implemented at several WWTPs (WERF 2009). Preliminary findings based on monitoring conducted at
nitrogen removal plants were presented at the WEF Specialty Conference: Nitrogen Removal 2009 and
are summarized below (Ahn et al 2009; WERF 2009).
• N2O emissions are related to an imbalance of the metabolic pathway or, in simpler terms,
recovery from stress. N2O can be produced by nitrifying bacteria in the aerobic and denitrifying
bacteria in the anoxic zone; however, denitrifying bacteria can consume N2O whereas nitrifying
bacteria cannot.
Nutrient Control Design Manual 2-17 August 2010
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• Measured N2O emissions from the aerobic zone were always higher than emissions from the
anoxic zone, contrary to previous thinking.
• N2O emissions from the aerobic zone exhibited spatial variability within the zone. The highest
N2O levels were observed close to the effluent.
• N2O emissions form a significant diurnal pattern that can be correlated with diurnal loading
patterns of ammonia.
• Release of N2O in the aerobic zone is largely a function of DO and ammonia concentrations.
Sampling campaigns at additional treatment plants are underway. Researchers are planning to publish
an interim Phase 1 report on the findings by the end of 2009 (WERF 2009).
2.10 References
Ahn, J.H., S. Kim, H. Park, D. Katehis, K. Pagilla, and K. Chandran. Spatial and Temporal Variability in N20
Generation and Emission from Wastewater Treatment Facilities. Presented at Nutrient Removal 2009.
Washington, DC. WEF.
Alexander, R.B., R.A. Smith, G.E. Schwarz, E.W. Boyer, J.V. Nolan, and J.W. Brakebill. 2008. Differences in
Phosphorus and Nitrogen Delivery to the Gulf of Mexico from the Mississippi River Basin. Environmental
Science and Technology. 42(3): 822-830. Available online:
http://water.usgs.gov/nawqa/sparrow/gulf findings.
Beline, F., J. Martinez, C. Marol, and G. Guiraud. 2001. Application of 15N Technique to Determine the
Contributions of Nitrification and Denitrification to the Flux of Nitrous Oxide from Aerated Pig Slurry.
Water Research. 35:2774-2778.
Bricker, S., B. Longstaff, W. Dennison, A. Jones, K. Boicourt, C. Wicks, and J. Woerner. 2007. Effects of
Nutrient Enrichment in the Nation's Estuaries: A Decade of Change. NOAA Coastal Ocean Program
Decision Analysis Series No. 26. Silver Spring, MD: National Centers for Coastal Ocean Science. 328 pp.
Available online: http://ccma.nos.noaa.gov/publications/eutroupdate/
Chesapeake Bay Program, 2008. Chesapeake Bay Program-A Watershed Partnership. Accessed July 1,
2008. Available online: http://www.chesapeakebay.net/nutrl.htm
Crites R. and G. Tchobanoglous. 1998. Small and Decentralized Wastewater Management Systems. New
York, NY: McGraw Hill.
EPA Region 10. 2007. Advanced Wastewater Treatment to Achieve Low Concentration of Phosphorus.
EPA Region 10. EPA 910-R-07-002.
Garrido J. M., J. Moreno, R. Mendez-Pampin, J.M. Lema. 1998. Nitrous Oxide Production under Toxic
Conditions in a Denitrifying Anoxic Filter. Water Research. 32(8):2550-2552.
Nutrient Control Design Manual 2-18 August 2010
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Gross, CM., J.A. Delgado, S.P. McKinney, H. Lai, H. Cover, and M.J. Shaffer. 2008. Nitrogen Trading Tool to
Facilitate Water Quality Trading. Journal of Soil and Water Conservation. March/April 2008. 63(2): 44-
45.
Hong, Z., Hanaki, K., and Matsuo, T. 1993. Greenhouse Gas-N2O Production during Denitrification in
Wastewater Treatment. Journal of Water Science Technology. 28(7): 203-207.
Hynes, R.K. and Knowles, R. 1984. Production of Nitrous Oxide by Nitrosomonas Europaea: Effects of
Acetylene, pH, and Oxygen. Canadian Journal of Microbiology. 30: 1397-1404.
Lo, I., K.V. Lo, D. Mavinic, D. Shiskowski, and W. Remay. 2008. Reduction of Nitrous Oxide Formation
from a Biological Nutrient Removal System in Wastewater Treatment- Factors of Aeration, External
Carbon and pH. In WEFTEC2008.
Kalogo, Y., and H. Monteith. 2008. State of Science Report: Energy and Resource Recovery from Sludge.
Prepared for Global Water Research Coalition, by WERF, STOWA, and UK Water Industry Research
Limited.
Kang, S.J., K.P. Olmstead, K.M. Takacs, and J. Collins. 2009. Sustainability of Full-Scale Nutrient Removal
Technologies. Presented at Nutrient Removal 2009. Washington, DC. WEF.
Kuai, L. and W. Verstraete. 1998. Ammonium Removal by the Oxygen-limited Autotrophic Nitrification-
Denitrification System. Applied Environmental Microbiology. 64: 4500-4506.
Landers, Jay. 2008. Halting Hypoxia. Civil Engineering. PP. 54-65. Reston, VA: ASCE Publications.
Long Island Sound Study. 2004. Protection+ Progress: Long Island Sound Study Biennial Report 2003-
2004. Project Manager/Writer Robert Burg, NEIWPCC/LISS. U.S. EPA Long Island Sound Office,
Stamford Government Center. Stamford, CT. Available online:
http://www.longislandsoundstudv.net/pubs/reports/30350report.pdf
Meyer, R.L., Zeng, R.J., Giugliano, V. and Blackall, L.L. 2005. Challenges for Simultaneous Nitrification,
Denitrification, and Phosphorus Removal in Microbial Aggregates: Mass Transfer Limitation and Nitrous
Oxide Production. FEMS Microbiology Ecology. 52:329-338.
Pehlivanoglu-Mantas, E. and D.L. Sedlak. 2006. Wastewater-Derived Dissolved Organic Nitrogen:
Analytical Methods, Characterization, and Effects - A Review. Critical Reviews in Environmental Science
and Technology. 36:261-285.
Poff, L.N., M. Brinson, and J. Day, Jr. 2002. Aquatic Ecosystems and Global Climate Change- Potential
Impacts on Inland Freshwater and Coastal Wetland Ecosystems in the United States. Prepared for the
Pew Center on Global Climate Change. January 2002.
Shiskowski, D.M. and Mavinic, D.S. 2006. The Influence of Nitrite and pH (nitrous acid) on Aerobic-Phase,
Auotrophic N2O Generation In A Wastewater Treatment Bioreactor. Journal of Environmental
Engineering and Science. 5: 273-283.
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Specter, M. 1998. Production and Decomposition of Nitrous Oxide during Biological Denitrification.
Water Environment Research. 70(5): 1096-1098.
STAC-WERF. 2007. Workshop Considerations and Presentations. Establishing a Research Agenda for
Assessing the Bioavailability of Wastewater-Derived Organic Nitrogen in Treatment Systems and
Receiving Waters, Baltimore, MD, September, 28, 2007.
State-EPA Nutrient Innovations Task Group. 2009. An Urgent Call to Action - Report of the State-EPA
Nutrient Innovations Task Group. Available online at
http://www.epa.gov/waterscience/criteria/nutrient/
Tallec, G., J. Gamier, and M. Gousailles. 2006. Nitrous Removal in a Wastewater Treatment Plant through
Biofilters: Nitrous Oxide Emissions during Nitrification and Denitrification. Bioprocess and Biosystem
Engineering. 29: 323-333.
Tallec, G., J. Gamier, G. Billen, and M. Gousailles. 2006. Nitrous Oxide Emissions from Secondary
Activated Sludge in Nitrifying Conditions of Urban Wastewater Treatment Plants: Effect of Oxygenation
Level. Water Research. 40: 2972 - 2980.
Tchobanoglous, G., F. L. Burton, and H.D. Stensel. 2003. Wastewater Engineering: Treatment and Reuse.
New York, NY: McGraw-Hill.
Third, K.A., N. Burnett, and R. Cord-Ruwisch. 2003. Simultaneous Nitrification and Denitrification Using
Stored Substrate (PHB) as the Electron Donor in an SBR. Biotechnology and Bioengineering. 83: 706-
720.
Thorn, M. and Sorensson, F. 1996. Variation of Nitrous Oxide Formation in the Denitrification Basin in a
Wastewater Treatment Plant with Nitrogen Removal. Water Resource. 30(6): 1543-1547.
USEPA. 1976. Process Design Manual for Phosphorus Removal. Great Lakes National Program Office.
GLNPO Library. EPA 625/1-76-OOla. April 1976.
USEPA. 1987. Handbook: Retrofitting POTWsfor Phosphorus Removal in the Chesapeake Bay Drainage
Basin. Center for Environmental Research Information. Cincinnati, OH. EPA/'625/6-87/017.
USEPA. 1993. Nitrogen Control Manual. Office of Research and Development. EPA/625/R-93/010.
September 1993.
USEPA. 1998. National Strategy for the Development of Regional Nutrient Criteria. Office of Water. EPA
822-R-98-002.
USEPA. 1999a. Wastewater Technology Fact Sheet: Fine Bubble Aeration. EPA 831-F-99-065.Available
online: http://epa.gov/OWM/mtb/mtbfact.htm
USEPA. 1999b. Wastewater Technology Fact Sheet: Sequencing Batch Reactors. EPA 832-F-99-073.
Available online: http://www.epa.gov/owm/mtb/sbr new.pdf
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USEPA. 2003a. Watershed-Based National Pollutant Discharge Elimination System (NPDES) Permitting
Implementation Guidance. EPA 833-B-03-004. Available online:
http://www.epa.gov/npdes/pubs/watershedpermitting finalguidance.pdf
USEPA. 2003b. Water Quality Trading Policy. Office of Water. Available online:
http://www.epa.gov/owow/watershed/trading/finalpolicy2003.pdf
USEPA. 2006. Wastewater Management Fact Sheet: Energy Conservation. Office of Water, Washington,
DC. EPA 832-F-06-024. http://www.epa.gov/owm/mtb/energycon_fasht_final.pdf
USEPA. 2007a. Opportunities for and Benefits of Combined Heat and Power at Wastewater Treatment
Facilities. U.S. Environmental Protection Agency, Combined Heat and Power Partnership, Washington,
DC. http://www.epa.gov/chp/documents/wwtf opportunities.pdf
USEPA. 2007b. Water Quality Criteria for Nitrogen and Phosphorus Pollution. Database Tools. Website
updated 27 September 2007. http://www.epa.gov/waterscience/criteria/nutrient/database/index.html
USEPA. 2007c. Watershed-based National Pollutant Discharge Elimination System (NPDES) Permitting
Technical Guidance. Office of Wastewater Management, Water Permits Division. Available online:
http://www.epa.gov/npdes/pubs/watershed techguidance entire.pdf
USEPA. 2008a. Emerging Technologies for Wastewater Treatment and In-Plant Wet Weather
Management. EPA 832-R-06-006. Available online: http://www.epa.gov/OW-
OWM.html/mtb/emerging technologies.pdf
USEPA. 2008b. Final Climate Change Strategy. Office of Water. Available online:
http://www.epa.gov/water/climatechange/strategy.html
USEPA. 2008c. Mississippi River Basin & Gulf of Mexico Hypoxia. EPA Office of Wetlands, Oceans and
Watersheds. Updated June 26, 2008. Available online: http://www.epa.gov/msbasin/
USEPA. 2008d. Municipal Nutrient Removal Technologies Reference Document., Volume 1 -Technical
Report. Office of Wastewater Management, Municipal Support Division. EPA 832-R-08-006. Available
online: http://www.epa.gov/OWM/mtb/mnrt-volumel.pdf
USEPA. 2008e. Online Technical Support for States: N-STEPS. Health and Ecological Criteria Division,
Office of Water Nutrient Criteria Program. Website updated 21 July 2008.
http://www.epa.gov/waterscience/criteria/nutrient/n-steps.html
USEPA. 2008f. Overview of Impaired Waters and Total Maximum Daily Loads Program. Wetlands, Oceans
& Watersheds. Website updated 16 October 2008. http://www.epa.gov/OWOW/TMDL/intro.html
USEPA. 2008g. State Adoption of Numeric Nutrient Standards (1998-2008). Office of Water. EPA821-F-
08-007. Available online: http://www.epa.gov/waterscience/criteria/nutrient/files/reportl998-2008.pdf
USEPA. 2008h. Water Quality Trading. Office of Oceans, Wetlands & Watersheds. Website updated 15
December 2008. http://www.epa.gov/owow/watershed/trading.htm
Nutrient Control Design Manual 2-21 August 2010
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USEPA. 2008L Ensuring a Sustainable Future: An Energy Management Guidebook for Wastewater and
Water Utilities. January 288. Available online:
http://www.epa.gov/waterinfrastructure/pdfs/guidebook si energymanagement.pdf
USEPA. 2009a. Methods for Evaluating Wetland Conditions. Website updated 9 February 2009.
http://www.epa.gov/waterscience/criteria/nutrient/guidance/wetlands/tfmodules
USEPA. 2009b. Nutrient Control Design Manual State of Technology Review Report. Office of Research
and Development. EPA/600/R-09/012. Available online:
http://www.epa.gov/nrmrl/pubs/600r09012/600r09012.pdf
U.S. Public Health Service and USEPA. 2008. Clean Watersheds Needs Surveys 2004 Report to Congress.
Available online: http://www.epa.gov/cwns/2004rtc/cwns2004rtc.pdf
Vanderploeg, H. 2002. The Zebra Mussel Connection: Nuisance Algal Blooms, Lake Erie Anoxia, and other
Water Quality Problems in the Great Lakes. 2002. Great Lake Environmental Research Laboratory. Ann
Arbor, Ml. Revised September 2002. Available online:
http://www.glerl.noaa.gov/pubs/brochures/mcvstisflver/mcystis.html
WEF and ASCE. 2006. Biological Nutrient Removal (BNR) Operation in Wastewater Treatment Plants -
MOP 29. Water Environment Federation and the American Society of Civil Engineers. Alexandria, VA:
WEFPress.
WERF. 2007a. Nutrient Challenge Research Plan - 2007. October 31, 2007. Available online:
http://www.werfnutrientchallenge.com/docs/Nutrient%20Removal%20Challenge%20Research%20Plan
%202007.pdf
WERF. 2007b. Nutrient Research Stakeholder Workshop. March 7 and 8, 2007. Baltimore, MD.
WERF. 2009. Molecular Level Through Whole Reactor Level Characterization of Greenhouse Nitrogen
Emission from Wastewater Treatment Operations. Project Number: U4R07. Updated: July 16, 2009.
Alexandria, VA: WERF.
Zeng, R.J., Lemaire, R., Yuan, Z. and Keller, J. 2003. Simultaneous Nitrification, Denitrification, and
Phosphorus Removal in a Lab-Scale Sequencing Batch Reactor. Biotechnology and Bioengineering.
84(2): 170-178.
Zheng, H., Hanaki, K., and Matsuo, T. 1994. Production of Nitrous Oxide Gas during Nitrification of
Wastewater. Water Science and Technology. 30(6): 133-141.
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3. Principles of Phosphorus Removal by Chemical Addition
Chapter 3 covers:
3.1 Introduction
3.2 Available Forms of Metal Salts and Lime
3.3 Equations and Stoichiometry
3.4 Solids Separation Processes
3.5 Effects on Sludge Production and Handling
3.6 Two Factors that May Limit the Ability of Plants to
Achieve Very Low Effluent Levels
3.7 References
3.1 Introduction
Chemical precipitation is widely used to remove phosphorus at wastewater treatment plants
(WWTPs). Chemicals such as metal salts or lime can be added to primary, secondary, or tertiary
processes or at multiple locations within the plant. Solids removal following chemical precipitation can
be conventional, such as primary or secondary clarification, or advanced, such as tertiary filtration or
alternative technologies. Chemical precipitation can be used alone or in conjunction with biological
phosphorus removal (BPR) to reduce the costs associated with the chemical dose and sludge production.
This chapter describes the principles of phosphorus removal by chemical addition. It includes a
description of chemicals available, the latest research on the mechanisms of phosphorus removal, and
factors affecting performance such as point of application and solids separation. Lastly, it presents
information on sludge production and limits of removal. Chapter 9 follows with more detailed
recommendations for designing a chemical phosphorus removal treatment system.
3.2 Available Forms of Metal Salts and Lime
Chemicals used for phosphorus precipitation are typically either metal salts or lime. The two
most common metal salts are aluminum sulfate (commonly known as alum) and ferric chloride. Sodium
aluminate can serve as a source of alum, although it can increase the pH substantially (WEF and ASCE
2009). Various forms of polyaluminum chloride (PAC) can also be used for chemical precipitation.
Ferrous sulfate and ferrous chloride, which are available as byproducts of steel-making operations
(pickle liquor), are also used. Lime is typically available in solid form as either quicklime (CaO) or
hydrated lime Ca(OH)2. Table 3-1 summarizes the most common chemical precipitants used for
phosphorus removal. Chapter 9 provides additional information on chemical properties and guidance
for chemical selection.
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Table 3-1. Common Chemicals Used for Phosphorus Removal
Chemical
Aluminum Sulfate
(Alum)
Sodium Aluminate
Polyaluminum
Chloride (PAC)
Ferric Chloride1
Pickle liquor (Ferrous
sulfate or Ferrous
iron)
Lime
Formula
AI2(S04)314(H20)
Na2AI2O4
AlnCI(3n-m)(OH)m
EX: AI12CI12(OH)24
FeCI3
Fe2SO4 or Fe2+
CaO, Ca(OH)2
Description
Crystallized, dry form or
liquid
Powder or liquid form.
Granular tihydrate is
common commercial form
Range in the degree of
basicity and aluminum
concentration
Orange-brown aqueous
solution
Green aqueous solution
Dry white powder or liquid
form as quicklime, CaO, or
hydrated lime, Ca(OH)2
Typical Weight
Percent in Commercial
Solutions2
48%
20%
51%
37 - 47%
Varies
NA
1. "Ferric is also common trade name for FeCIS and also 40% FeCI3 solution
2. Source: WEF and ASCE 2006 Tables 8.6 and 8.9
3.3 Equations and Stoichiometry
This section describes the forms of phosphorus that can be removed by chemical treatment. It
follows with a detailed discussion of the chemical reactions among metal salts, lime, and phosphorus
including factors affecting removal efficiency.
3.3.1 Removable Phosphorus
Chemical precipitation will remove only the phosphate (i.e., orthophosphate) fraction of total
phosphorus in wastewater. Influent phosphate is typically 50 to 80 percent of total phosphorus and
generally exists in one of two forms, H2PO4" and HPO42", with the first being dominant at pH below 8.3.
Polyphosphates will not react with metal salts or lime; however, they will be converted to phosphate
during biological treatment1. Organically bound phosphorus typically makes up the smallest fraction of
total influent phosphorus (1 < mg/L). The colloidal and particulate portion will generally be removed
during solids separation processes. The soluble organic fraction may either be hydrolyzed into
orthophosphate during the treatment process (if biodegradable) or will pass through a WWTP (if non-
biodegradable).
See Chapter 2, Section 2.2.2 for additional information on the forms of phosphorus in influent
wastewater. See Section 3.8 for a discussion on the implications of recalcitrant phosphorus on achieving
low effluent total phosphorus (TP) concentrations.
1 Colorimetric techniques used to quantify phosphate concentrations give results for "reactive" phosphorus, which
is primarily orthophosphate but includes a small condensed phosphate fraction.
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3.3.2 Reactions of Metal Salts and Phosphorus
The metal salt dose for chemical phosphorus removal has been recognized as a critical design
and economic parameter for chemical treatment systems. In addition to dose, the wastewater
characteristics, method of chemical addition, chemical addition feeding point(s), reaction pH,
flocculation method, and time after chemical addition are important design and performance issues that
can affect the relationship between dose and phosphorus removal efficiency.
The metal salt dose is commonly described in terms of the moles of metal added (Med0se) per
mole of soluble phosphorus in the influent (Pini). The term "stoichiometric dose" is based on 1.0 Me
added per Mole of P removed (i.e., 1.0 Medose/Pmi) which would be the molar ratio requirement for
strictly a chemical reaction between aluminum or iron salts to form a M-PO4 precipitate as shown in
Equations 3-1 and 3-2 (WEF and ASCE 2009, charges omitted).
AI2(SO4)3-14H2O+ 2H3(PO4) -» 2AI(PO4) + 3H2SO4 + 18H2O Eq. 3-1
-» FePo4+ 3CI + 2CO2+ 8 H2O Eq. 3-2
At relatively high effluent P concentrations (above 1.0 mg/L), the stoichiometric relationship for
metal salt dose is commonly observed. The molar ratio typically increases well above the stoichiometric
ratio as lower effluent phosphorus concentrations are needed. This has commonly been attributed to
substantial metal hydroxide formation in addition to metal-phosphorus precipitates (Sedlak, 1991).
Research by Szabo et al. (2008) and Smith et al. (2008) defined a mechanism for phosphorus
removal by ferric chloride based on a surface complexation model. Conceptually, the addition of ferric
chloride results in the precipitation of hydrous ferric oxide (HFO), which provides surface sites for
reaction with phosphorus. The basis of the phosphorus removal mechanism is that phosphate and iron
can share an oxygen molecule and that interactions can be represented by the following symbolic
reaction (charges omitted) (Smith et al. 2008):
FeOOH + HOPO3 = FeOOPO3 + H2O Eq. 3-3
The reactive oxygens are termed "surface sites," and their availability is related to mixing and aging
conditions. Under rapid mixing, surface sites are readily available. With slow mixing, however, much of
the HFO would form in the absence of phosphorus and result in less efficient phosphorus removal.
As the HFO forms, phosphate is simultaneously removed through (1) co-precipitation of
phosphate into the HFO structure, and (2) adsorption of the phosphate onto pre-formed HFO particles
(Smith et al. 2008). After the initial HFO formation and phosphorus removal, additional phosphorus can
be removed over time by diffusion of phosphorus into the floe. The effect of aging is to reduce reactive
sites and HFO phosphorus removal capacity. Additional research is needed to confirm a similar removal
mechanism for aluminum.
Chemical Dose and Phosphorus Removal Efficiency
In all observations on chemical removal of phosphorus, the percent phosphorus removal
increases and effluent phosphorus concentration decreases as the molar chemical dose for metal salts
increases, but the incremental removal diminishes with increased dosages (Sedlak 1991, Szabo et al.
Nutrient Control Design Manual 3-3 August 2010
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2008, WEF and ASCE 2009). Smith et al. (2007) found that for typical influent phosphorus
concentrations, ferric doses above 1.5 to 2.0 Medose/Pini ratios are sufficient to remove 80 to 98 percent
of soluble phosphorus. Reaching very low effluent TP concentrations, i.e. below 0.10 mg/L, requires
significantly higher ratios of about 6 or 7 Med0se/Pini • A similar dose was reported for 75 to 95 percent
phosphorus removal from wastewater using alum (WEF and ASCE 2009). For similar phosphorus removal
efficiencies using pre-polymerized salts such as PACI or sodium aluminate, a higher dose is needed.
Factors that influence dose and removal efficiency may include pH, alkalinity, competing substances in
the wastewater, initial mixing conditions and flocculation.
Mixing at the Dosage Point
Mixing at the dosage point is necessary to ensure that the metal and phosphate molecules
react. Mixing intensity can be represented by the velocity gradient, G, in units of seconds"1. Smith et al.
(2007) reported that the reactions of phosphates at the surface of the hydrous ferric oxides depend
highly on the mixing conditions. Rapid mixing means that the surface sites are available, whereas with
slower mixing, many of the metal oxides would form in the absence of phosphate and render internal
oxygen atoms unavailable for binding.
Bench-scale kinetic experiments by Szabo et al. (2008) revealed that a majority of phosphate will
react with iron in the first 10 to 20 seconds under ideal mixing conditions (G = 425 second"1). At
wastewater plants, mixing at the dosage point is typically poor, with G values ranging from 20 to 100
second"1 (Szabo et al. 2008). Szabo et al. recommend that plants attempt to achieve a very high mixing
intensity at the dosage point (G values between 200 and 300 second"1). Mixing times at high energy are
generally in the range of 10 to 30 seconds.
After the initial rapid kinetics, phosphorus removal can continue with slow reaction kinetics
between the phosphate and iron over many hours and even days (Szabo et al. 2008). The slow reaction
rate removal is more important for alum or ferric addition to activated sludge basins with solids
retention times in days.
Flocculation
After rapid mixing at the dosage point, gentle mixing is needed to form floes that can be settled
or removed through a solids separation process. This is critical for meeting low effluent phosphorus
requirements. Often, movement of the wastewater through the treatment plant is sufficient for floe
formation (USEPA 2008b). Flocculation can be limited by insufficient time or conditions that disrupt floe
formation such as pumping and aeration.
pH and Alkalinity
The highest removal efficiency for chemical precipitation is within a pH range of 5.5 to 7.0
(Szabo et al. 2008). Between pH 7 and 10, phosphorus removal efficiency declines because the surfaces
of metal hydroxides are more negatively charged, and soluble iron hydroxides begin to form. At low pH
values, the solubility of the precipitant is reduced, and at extremely low pH values, metal hydroxide
precipitation is limited. Szabo et al. (2008) reported similar relationships between phosphorus removal
efficiency and pH for both alum and ferric chloride.
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COD and TSS in Raw Wastewater
The efficiency of metal salt addition in the primary treatment step can be affected by the
wastewater characteristics. Based on the results of jar tests using municipal wastewater, Szabo et al.
(2008) observed a relationship between organic content and efficiency of phosphorus removal using
metal salts. Between chemical oxygen demand (COD) values of 300 to 700 mg/L, the phosphate
removal efficiency decreased with increasing COD. Similar results were observed for total suspended
solids (TSS), with lower phosphorus removal efficiency at higher TSS concentrations. In addition to
reduced efficiency during primary treatment, organic content can reduce the efficiency of metal salt
removal of phosphorus in activated sludge reactors. Iron and aluminum ions can react with humic and
fulvic acid substances to form insoluble complexes with the metal ions and their mineral oxides, thereby
blocking the reactive sites for phosphate precipitation (WERF 2009).
3.3.3 Reactions of Lime with Phosphorus
When lime is added to wastewater, it first reacts with the bicarbonate alkalinity to form calcium
carbonate (CaCO3). As the pH increases to more than 10, excess calcium ions will react with phosphate
to precipitate hydroxylapatite [Ca5(OH)(PO4)3] as shown in the reaction below.
5Ca2+ + 4OH" + 3HPO4. -^ Ca5OH(PO4)3 + 3H2O Eq. 3-4
Because it reacts first with alkalinity, the lime dose is essentially independent of the influent phosphorus
concentration. Tchobanoglous et al. (2003) estimates the lime dose to typically be 1.4 to 1.6 times the
total alkalinity expressed as CaCO3.
Lime addition can raise the pH to greater than 11. Because activated sludge processes require
pH levels below 9, lime cannot be added directly to biological treatment processes.
3.4 Solids Separation Processes
Solids separation methods are critical for determining phosphorus removal efficiency with
chemical precipitation as a large portion of the effluent phosphorus is contained in chemical
precipitates. Solids separation technologies, such as clarification and filtration are often used in
combination to achieve low effluent TP levels. Polymers can be used in addition to the metal salts for
phosphorus precipitation to enhance removal for fine particles and colloids.
Gravity separation in primary or secondary clarifiers is a traditional solids separation method at
WWTPs. Clarifiers used in chemical precipitation systems differ very little from those employed in
conventional treatment, although use of flocculation zones is recommended to provide flocculation time
after chemical addition.
For secondary clarification, flocculation can occur in aeration basins or channels preceding
clarification. The use of flocculation zones in secondary clarifiers is a recommended practice to allow
flexibility in the point of chemical addition and to provide a zone in which direct control can be exercised
over velocity gradients to achieve optimum flocculation.
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Membrane bioreactors for activated sludge wastewater treatment provide maximum solids
separation with effluent turbidity values typically less than 0.30 NTU and non-detect effluent TSS
concentrations.
Tertiary filtration following secondary clarification is increasingly being used as a polishing step,
with chemical to reduce TSS and to achieve effluent TP concentrations below 0.50 mg/L The types of
filters available include traditional media filters, upflow continuous backwash filters, cloth filters, and
membranes. Several patented ballasted high-rate clarifiers (BHRC) using different types of ballast such
as recycled sludge, microsand, and magnetic ballast (USEPA 2008b) have been developed in recent
years. Examples include DensaDeg®, Actiflo®, and the CoMag™ process. Chapter 6 of this design
manual provides an overview of the different tertiary filter technologies and Chapter 11 provides
recommendations for design.
3.5 Effects on Sludge Production and Handling
Sludge production and handling is generally considered to be one of the downsides of chemical
phosphorus removal. Chemical precipitation methods always produce additional solids due to
generation of metal- or calcium- phosphate precipitates and metal hydroxide sludge. The amount of
increased sludge production will depend on the location of chemical addition, the chemical dose used,
and the constituents present in the wastewater.
The stoichiometric relationships shown below can provide a good first estimate of additional
solids production from chemical precipitation (WEF and ASCE 2009). For alum addition the removal of P
with Al can be represented by Alo.sC^PCvXOH)!^ and the remaining aluminum added will be described
by aluminum hydroxide production in accordance with Eq. 3-5. For ferric removal with iron the sludge
production from P removal can be estimated as Fei.6(H2PO4)(OH)3.8 and the remaining ferric added will
be described by ferric hydroxide production in Eq. 3-6.
AI3+ + 3H2O <-> AI(OH)3 + 3H+ Eq. 3-5
Fe3+ + 3H2O <-> Fe(OH)3 + 3H+ Eq. 3-6
Typically, the addition of metal salts to the primary clarifier for the purposes of complete phosphorus
removal will increase primary treatment sludge production by 50 to 100 percent due to phosphorus and
hydroxide precipitates and increased suspended solids removal. In this case the secondary sludge
production is lower due to removal of additional TSS and BOD in the primary clarifier. The total overall
plant sludge production can be expected to increase by 60 to 70 percent (WEF and ASCE 2009). For
metal addition to the secondary treatment process to achieve effluent P concentrations in the range of
0.50 to 1.0 mg/L with the stoichiometric metal salt dose in the range of 2.0, the sludge production may
increase by 35 to 45 percent and the overall plant sludge production may increase by 5 to 25 percent
(WEF and ASCE 2009). For tertiary applications to achieve effluent P concentrations of less than 0.10
mg/L, the chemical stoichiometric dose can be 2 to 3 times that indicated in the previous sentence for
secondary treatment, but the amount of P to be removed is much less, so that the effect on sludge
production can be estimated to be increased by 45 to 60 percent for secondary/tertiary treatment and
by 10 to 40 percent for the overall plant sludge production. See Chapter 9, Section 9.8.3, for an example
calculation of sludge production increase resulting from the addition of metal salts.
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Lime typically produces a much higher sludge volume compared to metal salts because of its
reaction with natural alkalinity. The following reactions are important for determining the sludge
produced from lime addition (Tchobanoglous et al. 2003):
10Ca2+ + 6PO43" + 2OH" <-» Ca10(PO4)6(OH)2 Eq. 3-7
Mg2+ + 2OH" <->Mg(OH)2 Eq. 3-8
Ca2+ + CO32" <-» CaCO3 Eq. 3-9
An advantage of lime sludge is that some stabilization can occur due to the high pH levels required. One
disadvantage is that lime can cause scaling in mechanical thickening and dewatering systems. Although
alum tends to produce less sludge than ferric salts, alum sludge can be more difficult to concentrate and
dewater compared to ferric sludge.
The use of metal salts can result in increased inorganic salts (salinity) in the sludge and in the
effluent. Salinity can create problems when biosolids are land applied or when the effluent is returned
to existing water supply reservoirs. Biological phosphorus removal was developed in South Africa due to
the high rate of indirect recycling of wastewater effluent, which led to excessive total dissolved solids
(TDS) in the water supply during dry periods. High total salts can reduce germination rates for crops
and negatively affect the soil structure, in addition to potential taste and odor problems.
3.6 Two Factors that May Limit the Ability of Plants to Achieve Very Low Effluent Levels
Two factors may limit the ability of plants to achieve very low levels: the presence of
recalcitrant phosphorus and challenges in measuring very low effluent concentrations.
A portion of total phosphorus in wastewater can be recalcitrant, meaning that it passes through
the chemical, biological, and physical treatment processes and into the effluent unchanged. Lancaster
and Madden (2008) reported atypical recalcitrant phosphorus spikes as high as 0.5 mg/L in municipal
wastewater, which were suspected to originate from an industrial source. The authors speculated that
the recalcitrant fraction was made up of either dissolved acid-hydrolyzable phosphorus, organic
phosphorus, or a combination of the two. Neethling et al. (2007) postulated that recalcitrant
phosphorus is mostly the dissolved organic variety.
Studies conducted in Washington State and Nevada revealed significant variability in measured
phosphorus concentrations less than 0.020 mg/L. The City of Las Vegas Pilot Study found that the
accuracy of the measurement depended on several factors including the sample matrix and digestion
procedure (Eleuterio and Neethling 2009). In 2007, WERF initiated a project to evaluate the capabilities
of commercial laboratories to accurately measure phosphorus concentrations less than 0.020 mg/L and
determine factors that affect method accuracy. Key findings, which were presented at the WEF 2009
specialty conference on nutrient removal, are as follows:
• Total phosphorus measurements exhibited a wide variability. This range was attributed to
sample digestion procedures.
Nutrient Control Design Manual 3-7 August 2010
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• Orthophosphate measurements were accurate for de-ionized water, but not for wastewater
effluent samples. Variability in results was attributed to sample matrix and digestion
procedures.
• The Ascorbic Acid method (either the USEPA or Standard Methods procedure) was the most
often used and was found to be a reliable technique for measuring Orthophosphate at low
concentrations.
Eleuterio and Neethling (2009) concluded that additional research is needed to identify the compounds
interfering with the method and develop techniques to correct these interferences.
3.7 References
Benisch, M., D. Clark, Neethling, J.B., H.S. Fredrickson, A. Gu. 2007. Can Tertiary Phosphorus Removal
Reliably Produce 10 ug/L?: Pilot Results from Coeur D'Alene, ID. In Nutrient Removal 2007. WEF.
EPA Region 10. 2007. Advanced Wastewater Treatment to Achieve Low Concentration of Phosphorus.
EPA Region 10. EPA910-R-07-002.
Lancaster, C.D. and J.E. Madden. 2008. Not So Fast! The Impact of Recalcitrant Phosphorus on the
Ability to Meet Low Phosphorus Limits. In WEFTEC 2008.
Neethling, J.B., B. Bakke, M. Benisch, A. Gu, H. Stephens, H.D. Stensel, and R. Moore. 2005. Factors
Influencing the Reliability of Enhanced Biological Phosphorus Removal. Alexandria, VA: WERF and IWA
Publishing.
Neethling, J.B., M. Benisch, D. Clark, A. Gu. 2007. Phosphorus Speciation Provides Direction to Produce
10u.g/L \nNutrientRemoval2007. WEF.
Sedlak, R.I., editor. 1991. Phosphorus and Nitrogen Removal from Municipal Wastewater; Principles and
Practice, 2nd edition. Lewis Publishers, Boca Raton, Fl.
Smith, D.S., R.L. Gilmore, A. Szabo, I. Takacs, S. Murthy, and G. Daigger. 2008. Chemically Mediated
Phosphorus Removal to Low Levels: Analysis and Interpretation of Data. In WEFTEC2008.
Smith, S., A. Szabo, I. Takacs, S. Murthy, I. Licsko, and G. Daigger. 2007. The Significance of Chemical
Phosphorus Removal Theory for Engineering Practice. In Nutrient Removal 2007. WEF.
Smith, S., I. Takacs, S. Murthy, G.T. Daigger, and A. Szabo. 2008. Phosphate Complexation Model and Its
Implications for Chemical Phosphorus Removal. Water Environment Research. 80(5): 428-438.
Alexandria, VA: WEF.
Szabo, A., I. Takacs, S. Murthy, G.T. Daigger, I. Licsko, and S. Smith. 2008. Significance of Design and
Operational Variables in Chemical Phosphorus Removal. Water Environment Research. 80(5):407-
416. Alexandria, VA: WEF.
Tchobanoglous, G., F. L. Burton, and H.D. Stensel. 2003. Wastewater Engineering: Treatment and
Reuse. New York, NY: McGraw-Hill.
Nutrient Control Design Manual 3-8 August 2010
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USEPA. 2000. Wastewater Technology Fact Sheet: Chemical Precipitation. Office of Water. EPA832-F-
00-018.
USEPA. 2008a. Emerging Technologies for Wastewater Treatment and In-Plant Wet Weather
Management. EPA 832-R-06-006. Available online: http://www.epa.gov/OW-
OWM.html/mtb/emerging_technologies.pdf
USEPA. 2008b. Municipal Nutrient Removal Technologies Reference Document. Volume 1 -Technical
Report. Office of Wastewater Management, Municipal Support Division. EPA 832-R-08-006. Available
online: http://www.epa.gov/OWM/mtb/mnrt-volumel.pdf
WEF and ASCE. 1998. Design of Municipal Wastewater Treatment Plants - MOP 8, 4th Ed. Water
Environment Federation and American Society of Civil Engineers. Alexandria, VA: WEF.
WEF and ASCE. 2006. Biological Nutrient Removal (BNR) Operation in Wastewater Treatment Plants -
MOP 29. Water Environment Federation and the American Society of Civil Engineers. Alexandria, VA:
WEFPress.
WEF and ASCE. 2009. Design of Municipal Wastewater Treatment Plants - WEF Manual of Practice 8
and ASCE Manuals and Reports on Engineering Practice No. 76, 5th Ed. Water Environment
Federation, Alexandria, VA, and American Society of Civil Engineers Environment & Water Resources
Institute, Reston, Va.
WERF. 2009. Chapter 1, Introduction: Tertiary Phosphorus Removal. Nutrient Compendium. Website
accessed 11 March 2009. http://www.werfnutrientchallenge.com/chapterl.asp?area=chl
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4. Principles of Biological Nitrogen Removal
Chapter 4 covers:
4.1 Introduction
4.2 Nitrogen Removal by Biomass Synthesis
Nitrification
4.3 Microbiology of Nitrification
4.4 Reactions and Stoichiometry of Nitrification
4.5 Nitrification Kinetics
4.6 Inhibitory Effects of Environmental Conditions on Nitrification
Denitrification
4.7 Denitrification Fundamentals
4.8 Microbiology of Denitrification
4.9 Metabolism and Stoichiometry of Heterotrophic Denitrification
4.10 Biological Denitrification Kinetics with Influent Wastewater
4.11 Denitrification Carbon Sources and Relative Consumption Rates
4.12 Denitrification Kinetics of Exogenous Carbon Sources
4.13 Specific Denitrification Rates
Additional Topics
4.14 Simultaneous Nitrification-Denitrification
4.15 Metabolism, Stoichiometry, and Kinetics of ANAMMOX®
4.16 Impacts on Sludge Production and Handling
4.17 Effluent Dissolved Organic Nitrogen
4.18 References
4.1 Introduction
Biological nitrogen removal in wastewater treatment occurs by two primary mechanisms: 1)
biomass synthesis (nitrogen assimilation) and sludge wasting, and 2) biological nitrification and
denitrification, with only the latter able to achieve high levels of nitrogen removal and low effluent
concentrations of inorganic nitrogen in biological nutrient removal processes treating domestic
wastewaters. Nitrification is a two-step process in which one genus of aerobic bacteria oxidize
ammonia-nitrogen (NH3-N) to nitrite-nitrogen (NO2-N) followed by another genus which oxidizes nitrite-
nitrogen to nitrate-nitrogen (NO3-N). Under certain conditions, e.g. inadequate dissolved oxygen, the
process can be stopped at NO2-N formation. In biological denitrification, a carbon source is oxidized
using nitrate and/or nitrite as electron acceptors in biological oxidation-reduction reactions to reduce
the oxidized nitrogen (NO3-N or NO2-N) to inert nitrogen gas (N2). An anaerobic process that does not
require a carbon source for NO2-N reduction is the ANAMMOX® (anaerobic ammonia oxidation)
process in which certain bacteria are capable of oxidizing ammonia with nitrite reduction to produce N2
(Sliekersetal. 2002).
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This chapter provides an overview of the principles behind biological nitrogen removal including
autotrophic and heterotrophic microorganisms responsible for nitrification and denitrification, reactions
and stoichiometry, kinetics, effects of environmental factors, and fixed film process considerations, as
well as simple design calculations.
4.2 Nitrogen Removal by Biomass Synthesis
Biochemical oxygen demand (BOD) removal in biological wastewater treatment processes
occurs with bacterial cell growth. A commonly used formulation for biomass (Hoover and Porges 1952)
is C5H7O2N, which suggests that nitrogen accounts for 12 percent of the biomass volatile suspended
solids (VSS). For domestic wastewater treatment, 15 to 30 percent of influent nitrogen can be removed
via biomass synthesis and sludge wasting. If the waste sludge is processed by anaerobic digestion, about
one half of the nitrogen removed can be released as ammonia in the digester and returned to the
biological treatment process in dewatering recycle streams, such that only 10 to 15 percent net removal
by synthesis occurs. Centrate or filtrate from digester solids dewatering can contain NH4-N
concentrations ranging from 500 to 1200 milligrams per liter (mg/L), depending on the level of solids
thickening prior to anaerobic digestion and the digester operation.
The nitrogen removal efficiency by biomass synthesis depends on the biological process influent
BOD to total nitrogen (TN) ratio and the biomass solids retention time (SRT) in the system. With
increasing SRT, a greater fraction of the biomass undergoes endogenous decay, releasing NH4-N and
decreasing the net nitrogen removal due to both synthesis and decay. Not all of the nitrogen is released
with cell loss, as some of it remains unavailable in cell debris following cell death. The effect of SRT and
influent BOD/TN ratio on the percent nitrogen removal due to biomass synthesis and decay is illustrated
in Figure 4-1. At lower SRTs and higher influent BOD/N ratios, the removal due to biosynthesis and
sludge wasting is higher. Assumptions used were a biomass synthesis yield value of 0.50 g VSS/g BOD
removed and an endogenous decay rate of 0.08 gVSS/gVSS per day (Tchobanoglous et al. 2003). The
fraction of biomass debris remaining from endogenous decay was assumed to be 0.08 gVSS/gVSS
(Barker and Dold 1997).
Nutrient Control Design Manual 4-2 August 2010
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35
BOD/N=6.0
-o--BOD/N=4.0
5 10 15
SRT, days
Figure 4-1. Percent nitrogen removal due to biomass synthesis as a function of SRT and influent BOD/N ratio.
4.3 Microbiology of Nitrification
Ammonia- and nitrite-oxidizing bacteria are referred to as autotrophic bacteria because they use
carbon dioxide (CO2) for their carbon source. More specifically, they are aerobic chemoautotrophic
bacteria because, in addition to using CO2, they require dissolved oxygen to oxidize an inorganic
compound (NH4-N or NO2~N) to obtain cell energy. A key functional enzyme possessed by all these
bacteria is ammonia monooxygenase (AMO). This enzyme oxidizes ammonia to hydroxylamine, which is
subsequently converted to nitrite by hydroxylamine oxidoreductase. The ammonia-oxidizing bacteria
(AOB) are designated as Nitroso organisms and include the genera Nitrosomonas, Nitrosospira, and
Nitrosococcus. The nitrite-oxidizing bacteria (NOB) are designated as Nitro-organisms and include the
genera Nitrobacter, Nitrospira, Nitrococcus, and Nitrospina (Rowan et al. 2003).
Differences in 16S ribosomal RNA (rRNA) sequences for AOB and NOB genera results in distinct
differences in phylogenics (Rowan et al. 2003). The phylogenetic distribution of AOB is summarized in
Table 4-1. The genera Nitrosomonas and Nitrosospira are soil and freshwater AOB that are in the class
6-Proteobacteria. Five lineages of Nitrosomonas AOB have been determined by Koops and
Pommerening-Roser (2001) and are N. europaea/eutropha, N. communis, N. oligotropha, N. marina, and
N. cryotolerans. Nitrosococcus AOB are located in the a-Proteobacteri and consist of only marine AOB
species with strains Nitrosococcus oceani and Nitrosococcus halophilus. Nitrosococcus mobilis was
previously misnamed and now belongs to the 6-Proteobacteri genus Nitrosomonas and not the genus
Nitrosococcus in the a-Proteobacteria.
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Table 4-1. Phylogeny of Ammonia-Oxidizing Bacteria
Subclass of
Proteobacteria
6-
Proteobacteria
a-Proteobacteria
Sub Clusters
europaea-
mo bills
communis
oligotrophia
marina
nitrosospira
Species
Nitrosomonas
europea
Nitrosomonas
eutropha
Nitrosomonas
halophilar
Nitrosococcus mobilis
Nitrosomonas
communis
Nitrosomonas sp. 1
Nitrosomonas sp. II
Nitrosomonas nitrosa
Nitrosomonas ureae
Nitrosomonas
oligotropha
Nitrosomonas marina
Nitrosomonas sp. Ill
Nitrosomonas
aestuarii
Nitrosomonas
cryotolerans
Nitrosolobus
multiformis
Nitrosovibrio tenuis
Nitrosospira sp. 1
Nitrosococcus oceani
Salt
Requirement
Halotolerant,
moderately
halophilic
No salt requirement
No salt requirement
No salt requirement
Obligately halophilic
Obligately halophilic
No salt requirement
Obligately halophilic
NH3
affinity
Ks,uM
30-61
14-43
19-46
1.9-4.2
50-52
42-59
Preferred
Habitat
WWTP, eutrophic
freshwater, brackish
water
Soils (not acid)
Eutrophic
freshwater
Oligotrophic
freshwater,
natural soils
Marine environment
Soils (not acid)
Soils, rocks, and
freshwater
Marine environment
Source: Modified from Koops and Pommerening (2001).
The four NOB genera show more diverse phylogenetics with Nitrobacter within a-Proteobacteri,
Nitrococcus within y-Proteobacteria, Nitrospina within the S-Proteobacteria, and the two species of the
genus Nitrospira, Nitrospira moscoviensis and Nitrospira marina within a separate phylum close to the 6
subclass of the Proteobacteria (Siripong and Rittmann 2007). These are summarized in Table 4-2 (Koops
and Pommerening-Roser 2001). While Nitrobacter generally occur as free cells, Nitrospira are more
commonly observed attached to floes or biofilms in their natural environments.
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Table 4-2. Phylogeny of Nitrite-Oxidizing Bacteria
Subclass of
Proteobacteria
a-Proteobacteria
y-Proteobacteria
6-Proteobacteria
Species
Nitrobacter alkalicus
Nitrobacter winogradskyi
Nitrobacter vulgaris
Nitrobacter hamburgenis
Nitrococcus mobilis
Nitrospina gracilis
Nitrospira moscoviensis
Nitrospira marina
Ecophysiological
parameters
Alkali and halotolerant
No salt requirement
Obligately halophilic
Obligately halophilic
No salt requirement
Obligately halophilic
Preferred
Habitat
Soda lakes
Freshwater, soils, rocks
Marine environments
Freshwater
Marine environment
Source: Modified from Koops and Pommerening (2001).
Using molecular biology tools, researchers have identified diverse populations of AOB and NOB
in activated sludge and fixed film nitrification systems. Siripong and Rittmann (2007) examined the
diversity of nitrifying bacteria communities in the seven activated sludge facilities operated by the
Metropolitan Water Reclamation District of Greater Chicago (MWRDGC) with varying SRT, temperature,
and influent characteristics. They found the coexistence of AOB Nitrosomonas and Nitrosospira genera
in all facilities and the coexistence of NOB Nitrobacter and Nitrospira genera. Colder temperatures
appeared to increase proportions of Nitrosospira and Nitrospira. Park et al. (2002) found that the AOB
community in an aerated-anoxic process was very diverse, with both Nitrosomonas and Nitrosospira
present. Others have shown that Nitrosomonas is very common in activated sludge treatment and
Nitrospira is the most common NOB (Aoi et al. 2000; Coskuner and Curtis 2002; and Harms et al. 2003; Li
etal. 2005).
In a laboratory fixed film fluidized bed reactor, Schramm et al. (1999) found a dominance of
Nitrosospira for AOB and Nitrospira for NOB. In contrast, Rowan et al. (2003) found a dominance of
Nitrosomonas for AOB in a biological aerated filter and trickling filter receiving combined domestic and
industrial wastewater.
Though their role in wastewater treatment has not been determined, it is notable that the
diversity of ammonia oxidation via AMO extends to the domain Archaea with ammonia oxidation found
in members to the kingdom Crenarchaeota. Originally thought to be a marine organism with their
discovery by Fuhrman et al. (1992), they have also been found in a wide range of soils, in sediments, and
in freshwater (Nicol and Schleper 2006). Furthermore, Leininger et al. (2006) found that they were
more abundant than autotrophic bacteria for ammonia oxidation in soils. A marine Crenarchaeota
isolated by Konneke et al. (2005) was able to grow as a chemoautotroph at rates comparable to AOB,
and it also had a much higher affinity for ammonia (0.03 to 1.0 u.M). However, the presence of organic
substrate appeared to inhibit its growth.
4.4 Reactions and Stoichiometry of Nitrification
The energy yielding ammonia oxidation reaction by AOB is as follows:
NhLT + 1.5O -* NO + 2H+
+ H2O
Eq. 4-1
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In Equation 4-1, 1.5 moles of oxygen are required to oxidize 1 mole of ammonia and 2 moles of
hydrogen are produced. Thus, the oxygen requirement to produce nitrite is 3.43 g O2/g NH4-N oxidized
to NO2-N. The energy producing reaction by NOB is as follows and shows the need for 1.14 g O2/g NO2-
N oxidized to NO3-N:
N02+0.502^N03 Eq.4-2
The overall reaction for the two-step oxidation process can be written as follows and accounts
for alkalinity consumption by the hydrogen produced:
NH;+202+2HC03^N03+2C02+3H20 Eq. 4-3
Based on the above stoichiometry, ammonia oxidation to nitrate requires 4.57 g O2/g NH4-N
oxidized to NO3-N and consumes 7.14 g alkalinity (as CaCO3) per g NH4-N oxidized (Tchobanoglous et al.,
2003).
The oxygen required and alkalinity consumption per g NH4-N removed calculated from Equation
4-3 will be less in reactors because some of the NH4-N removed is consumed for biomass synthesis by
the nitrifying bacteria. Werzernak and Gannon (1967) found that the oxygen consumption normalized
to nitrogen removal was 4.33 g O2/g NH4-N oxidized to NO3-N with 3.22 g O2 used for NH4-N oxidation to
NO2-N and 1.11 g O2 for oxidation of NO2-N to NO3-N. This fits reasonably close to the stoichiometry
presented by Haug and McCarty (1972) in which biomass yields of 0.15 g VSS/g NH4-N and 0.02 g VSS/g
NO2-N removed were determined. Their balances are as follows:
Ammonia consumption:
55NH;+7602+109HC03 -»
Eq. 4-4
C5H7NO 2 + 54NO ' + 104H 2CO 3 + 57H 20
Nitrite consumption:
Eq. 4-5
400NO - + NH; + 1950 2 + HCO 3 + 4H2CO3 -»
C5H7NO2 + 400NO 3 + 3H2O
When accounting for synthesis per the above reactions, the calculated oxygen and alkalinity
consumptions for NH4-N removal are lower by about 5 percent and 1 percent, respectively. At longer
SRTs with decay of nitrifiers, ammonia will be released and the difference is less. By ignoring synthesis
and using the oxidation only values of 4.57 g Ojq NH4-N oxidized to NO3-N and 7.14 g alkalinity (as
CaCO3) perg NH4-N oxidized, the results are slightly conservative but practical for design estimates.
4.5 Nitrification Kinetics
Where ammonia removal is needed, nitrification kinetics will govern the activated sludge
aerobic zone design as the nitrifying bacteria have slower growth rates then the BOD-consuming
heterotrophic bacteria and thus require a longer SRT. Bacterial growth rate models (Monod model) or
Nutrient Control Design Manual 4-6 August 2010
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substrate utilization models are commonly used to describe nitrification kinetics. The models are used
to fit data from steady state or non-steady state nitrification systems to obtain appropriate coefficients.
Because it has a significant effect on AOB and NOB growth rates, substrate utilization rates, and
endogenous decay rates, temperature is accounted for in nitrification kinetic models. The models will be
reviewed first before presenting kinetic coefficient information.
The Monod model is used to describe the specific growth rate of nitrifying bacteria as a function
of the concentration of the nitrogen species being oxidized, DO concentration, and endogenous decay
rate as shown in Equations 4.6 and 4.7 below for AOB and NOB, respectively. For completely mixed
activated sludge systems at temperatures below 25°C, the process is sufficiently modeled by considering
only the AOB, as the NOB are able to use nitrite much faster. However, at higher temperatures, the
kinetics of both groups must be considered as the effect of higher temperature favors AOB more than
NOB. This is an underlying basis for the Single Reactor High Activity Ammonia Removal Over Nitrite
(SHARON®) process for nitrogen removal, which is described later in this chapter. For nitrification in
batch fed systems or for aerobic reactors in series, it may be more appropriate to model both AOB and
NOB to determine treatment NH4-Nand NO2-N concentrations (Chandran and Smets, 2000).
max,AOB
C i If
°NH """ ^N
S +K
o,AOB
-b.
Eq. 4-6
*NOB
*max,NOB
JNO
c , if
° ^'^
S +K
o,NOB
-b
NOB
Eq. 4-7
Where:
M-AOB
M-NOB
HMAX,AOB
HMAX,NOB
bAOB
bNOB
SNH
KNH
S0
KO,AOB
KO,
,NOB
= Specific growth rate of ammonia-oxidizing bacteria, g VSS/g VSS - d
= Specific growth rate of nitrite-oxidizing bacteria, g VSS/g VSS - d
= Maximum specific growth rate of ammonia-oxidizing bacteria, g VSS/g VSS - d
= Maximum specific growth rate of nitrite-oxidizing bacteria, g VSS/g VSS - d
= Specific endogenous decay rate of ammonia-oxidizing bacteria, g VSS lost/g VSS - d
= Specific endogenous decay rate of ammonia-oxidizing bacteria, g VSS lost/g VSS - d
= NH+4- N concentration, mg/L
= Half-velocity coefficient for NH+4- N, mg/L
= DO concentration, mg/L
= Half - velocity coefficient for DO for AOB, mg/L
= NO~2 - N concentration, mg/L
= Half - velocity coefficiency for NO"2 - N, mg/L
= Half - velocity coefficient for DO for NOB, mg/L
The next several equations describe the specific growth rate of AOB, volumetric NH4-N oxidation
rates, and effluent NH4-N concentration as a function of kinetic parameters and system SRT. An
identical set of equations can be used for NOB to describe NO2-N oxidation kinetics and are not written
here.
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*AOB
SRT
Eq. 4-8
KNH(l + bAOBSRT)
SRT(|a
max,AOB,DO "AOB
-bAOB)-1.0
Eq. 4-9
>max,AOB)(So)
max, AOB,DO
Eq. 4-10
Where:
SRT
Mrnax,AOB,DO
= Solids retention time, days
= M-max,AOB corrected for DO concentration, g/g - day
The next set of kinetic equations describes nitrification rates in batch tests using activated
sludge or nitrifier enrichments grown in bench-scale studies. In this case, only the equations for AOB are
shown but a similar set of equations would apply to NOB. From the Monod equation, bacteria-specific
growth rate can be described as a function of substrate utilization.
'x,AOB
I-1 AOB .. I"1
X
max,AOB
AOB
,SNH +KNH I 5o +Ko,AOB
Eq. 4-11
Where:
|"X,AOB = AOB growth rate, mg/L - day
XAOB = AOB concentration, mg/L
x,AOB r*max,AOB
C I If
°NH ~"~ IXNH
S +K
o,AOB
X
AOB
Eq. 4-12
The rate of biomass growth is a function of the substrate utilization rate and synthesis yield coefficient:
oB(rNH) Eq.4-13
x,AOB
Where:
rNH = NH+4 - N oxidation rate, mg/L - day
Combining Equations 4.12 and 4.13 gives the NH4-N utilization rate.
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max,AOB
+K
NH NH
S +K
X
AOB
Eq. 4-14
Besides the kinetic model parameters, the ammonia utilization rate depends on the AOB
concentration, which can be calculated from a mass balance on the average daily amount of ammonia
oxidized, the AOB synthesis yield and endogenous decay coefficient, and the average SRT
(Tchobanoglous et al. 2003).
XAOB ~
Q(YAOB)(Nox)SRT
BSRT)
Eq. 4-15
Where:
XAOB
Q
Nox
V
= AOB concentration, mg/L
= Average daily influent flow, L/day
= NH+4 - N oxidized by AOB from influent, mg/L
= Volume of reactor containing AOB, L
Combining Equations 4-14 and 4-15 gives the NH4-N utilization rate with the growth reactor AOB:
S +K
Q(Nox)SRT
BSRT)
Eq. 4-16
Equations 4-9 and 4-16 represent expressions that can be used for two different types of tests
to obtain nitrification kinetics. In the first case, the effluent NH4-N concentration is measured at
different operating SRTs to fit the model to obtain the maximum specific growth rate, u.max,AOB- For
Equation 4-16, careful measurement of NH4-N oxidation over time and SRT control and maintenance is
needed before using the reactor biomass in batch nitrification tests in which rNH versus SNH is measured
to obtain u.max,AOB by model fitting. It should be noted that in both cases the value used for the
endogenous decay rate, bAOB, affects the calculated u.max,AOB value.
4.5.1 AOB kinetics
In the nitrification model equations shown previously, the value used for the specific decay
coefficient directly affects the calculated u.max,AOB value. For a given observed nitrification rate or effluent
NH4-N concentration with SRT, higher specific decay coefficient values result in higher u.max,AOB values and
vice versa. Thus, it is difficult to compare nitrification kinetic values between studies without knowing
the value for bAOB, and the maximum specific growth rate is not accurately determined without an
accurate determination of the specific endogenous decay rate coefficient (Dold et al. 2005).
In early work on nitrification kinetics, the specific endogenous decay rate was believed to be
very low and generally ignored (Downing et al. 1964; USEPA 1993). Tests evaluating AOB nitrification
kinetics revealed the importance of the specific endogenous decay rate value (Melcer et al. 2003).
Similar u.max,AOB values were obtained in a study using three different experimental methods to obtain
nitrification kinetics and applying a bAOB value of 0.17 g/g-d obtained in a separate test method (Dold et
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al. 2002; Melcer et al. 2003). The specific endogenous decay coefficient obtained in the Water
Environment Research Foundation (WERF) study is included in Table 4-3 and compares closely to results
from other studies aimed at determining bAOB. Manser et al. (2006) obtained similar specific
endogenous decay rates for AOB and NOB and, in addition, obtained similar rates for membrane
bioreactor (MBR) and conventional clarification activated sludge.
Lower specific endogenous decay coefficients have been found for anoxic-aerobic systems with
Lee and Oleszkiewicz (2003), Nowak et al. (1994), and Siegrist et al. (1999) reporting about a 50-percent
reduction in the decay rate. Temperature is also known to affect the endogenous decay rate of both
heterotrophic and autotrophic microorganisms. The effect of temperature on the AOB endogenous
decay rate was also determined in the WERF study (Melcer et al. 2003) from 10 to 20°C, and the
temperature coefficient, 6, was found to be 1.029 in the temperature relationship given in Eq. 4-17
below.
'AOB,T
Eq. 4-17
Where:
bAOB,20
e
= Endogenous decay rate at temperature T, °C
= Endogenous decay rate at 20oC
= Temperature coefficient, 1.029
Table 4-3. Summary of Test Results on Measuring Specific Endogenous Decay Coefficient
Rates (All Rates at 20°C)
Bacteria or source
Aerobic activated sludge
Nitrobacter
Aerobic activated sludge
Aerobic activated sludge
Aerobic activated sludge
Aerobic activated sludge
Aerobic activated sludge
Aerobic activated sludge
Anoxic-aerobic activated sludge
Aerobic activated sludge
Nitrifier
group
AOB
NOB
AOB
AOB
NOB
AOB
AOB
AOB
AOB
AOB
Specific endogenous
decay rate (b), g VSS/g
VSS-day
0.17
0.14
0.17
0.15*
0.14*
0.20**
0.21**
0.15
0.10
0.09
Reference
Melcer etal., 2003
Copp and Murphy, 1995
Copp and Murphy, 1995
Manser etal., 2006
Manser etal., 2006
Nowak etal., 1994
Siegrist etal., 1999
Lee and Oleszkiewicz, 2002
Lee and Oleszkiewicz, 2003
Katehis etal., 2002
* Showed similar results for membrane bioreactor and conventional activated sludge systems.
** Reported that the decay rate under anoxic conditions was about 1/2 of the aerobic system rate.
Often referred to as U,AUT , the jumox value for AOB and NOB is a critical kinetic parameter and
shown to be a function of temperature. AOB has historically been the focus of these kinetic evaluations,
and results on the 20°C value and effect of temperature have varied widely. Reported values for early
studies on AOB u.max at 20°C range from 0.32-0.77 g/g-day (Downing et al. 1964; Downing and Hopwood
1964; Barnard 1975; Lawrence and Brown 1976; Hall and Murphy 1980; Randall etal. 1992), but these
studies assumed negligible or very low specific endogenous decay rates. At the higher measured decay
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rate, the WERF study (Melcer et al. 2003) obtained a 20°C u.max value of 0.90 g/g-day for three different
test methods. However, more recent studies also show a wide range in AOB 20°C u.max values; 0.63 -
4.55 g/g-day (Chandran et al. 2008; Marsili-Libelli et al. 2001, and lacopozzi et al. 2007).
Similarly, the reported effect of temperature on the 20°C u.max value has also ranged widely with
the ratio of the 20°C u.max /10°C u.max from 2.2 to 3.7. The ratio in the WERF study (Melcer et al. 2003) was
2.0. Reported differences in kinetic data are the result of data analysis methods, activated sludge
characteristics, and operating conditions. Comparing or using only single kinetic parameter values
between studies may not be valid as the calculated kinetic value is related to the determined or
assumed values for other kinetic parameters in the nitrification model. Because a complete set of
model parameters were determined in the evaluation of AOB kinetics in the WERF study (Melcer et al.
2003), these values are given in the following nitrification AOB kinetic values in Table 4-4 for use in
nitrification kinetic modeling (Eq. 4-9 through 4-16). These studies were done at DO concentrations in
the 4-6 mg/L range so the DO half-saturation value was not evaluated. The value from the IWA activated
sludge model report is used here (Henze et al. 2000). Designers should not take these values as
absolute and if at all possible measure and calibrate nitrification kinetic values for the design facility
(See Chapter 10, Section 10.8 for additional discussion).
Table 4-4. Summary of AOB Nitrification Kinetic Coefficient Values.
Parameter
Yield, Y
Specific endogenous decay
rate, b
Maximum specific growth
rate, |amax
Half-velocity coefficient,
KNH
Half-velocity coefficient, K0
Units
g VSS/g N oxidized
g VSS/g VSS-day
g VSS/g VSS-day
mg/L
mg/L
20°C
Value
0.15
0.17
0.90
0.70
0.50
Temperature
correction value, 6
1.0
1.029
1.072
1.0
1.0
Source: All from Melcer et al. (2003) except the DO half-velocity coefficient; from Henze et al. (2000).
Mixing, floe size, and site specific operating conditions can affect nitrification kinetics. Large floe
sizes and less mixing will have more diffusion limitations with lower inorganic nitrogen and DO
concentrations within the floe. Thus, lower specific nitrification rates will be observed as the rates are
normalized to the total VSS concentration or biomass concentration. The effect of kinetic limitations
was discussed by Manser et al. (2005) in which they observed lower AOB and NOB half-velocity
coefficient values (Table 4-5) for a conventional activated sludge versus MBR. The MBR had smaller floe
size, which they reasoned was due to greater agitation for membrane scour. The effect was greatest on
the oxygen half-velocity coefficient. If constant half-velocity coefficient values were assumed in
interpreting the test data, a lower observed u.max value would have been determined.
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Table 4-5. Comparison of Nitrification Half-Velocity Coefficients (mg/L) in MBR and
Conventional Activated Sludge (CAS) Systems
Nitrifier group
AOB
Oxygen, Ko
Ammonia, KNH
NOB
Oxygen, Ko
Ammonia, KNH
MBR
0.18
0.13
0.13
0.17
CAS
0.79
0.14
0.47
0.28
Source: Manser et al. 2005.
In addition to floe size, the oxygen uptake rate (OUR) of the activated sludge can affect the
observed nitrification rate at a given bulk liquid DO concentration due to more oxygen limitation in the
floe at greater OURs. Stensel et al. (1993) showed that for a staged nitrification aeration tank, a higher
DO concentration was needed to maintain the same nitrification rate at a given bulk liquid NH4-N
concentration for mixed liquor in the initial aerobic stages versus mid to latter stages. The higher OUR
resulted in greater oxygen limitations in the activated sludge floe and affected nitrification rates. The
effects of mixed liquor suspended solids (MLSS) concentration, floe size, and OUR are not currently
factored into nitrification kinetic models. Also note that the transfer rate of DO to the nitrifiers ceases
to be a significant nitrification factor in long hydraulic residence time (HRT) - long SRT systems such as
extended aeration systems. For these systems, complete nitrification can be accomplished at very low
DO concentrations, i.e. <0.3 mg/L, even though the kinetic rate would be faster if DO concentration was
higher.
4.5.2 NOB kinetics
NOB kinetics are of major interest with regard to treating high temperature anaerobic digestion
centrate or filtrate streams for nitrogen removal with the SHARON® process (discussed below) or for
operating conditions that lead to higher effluent NO2-N concentrations. Studies on NOB kinetics have
been based on model fitting of pilot plant or bench-scale results in which selected parameters have
been arbitrarily fixed or on evaluating one particular kinetic effect such as DO concentration. Previous
work has shown that NOB are inhibited at low DO concentrations with NO2-N concentrations increasing
at low DO (Picioreanu et al. 1997; Garrido et al. 1997; Peng and Zhu 2006; Contreras et al. 2008),
suggesting a higher K0 for NOB than AOB. Values for K0 for studies with NOB ranged from 0.50 to 1.75
mg/L. For studies with both NOB and AOB, the ratio of the NOB to AOB DO half velocity coefficients
were 2.36 (Guisasola et al. 2005), 1.4 (Ciudad et al. 2006), and 0.59 (Manser et al. 2005). Absolute
values of oxygen half-velocity coefficients depend on the operating and activated sludge floe conditions
and possible diffusion limitations in addition to the bacterial affinity for DO.
Nitrite oxidation Kinetic coefficients used in model fits vary widely (Sin et al., 2008). Lacopozzi
et al. (2007) presented a model for the two-step nitrification process using 20°C u.max values from Marsili-
Libellie et al. (2001) of 0.63 g/d-d for AOB and 1.04 g/g-d for NOB. Based on Eq. 4-17, their assumed
temperature coefficient, 6, values for u.max were 1.06 for both AOB and NOB. However, Kaelin et al.
(2009) fit a two-step AOB and NOB kinetic model to BNR pilot plant data at winter and summer
conditions with constant or intermittent aeration. Their data fit resulted in 20°C u.max values of 0.90 and
0.65 g/g-d for AOB and NOB, respectively. They found that the effect of temperature was more
pronounced with AOB versus NOB, and based on Eq. 4-17, their temperature coefficient, 6, values for
both u.max and b were 1.13 for AOB and 1.08 for NOB. These results are consistent with observations that
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at higher temperatures, above 25°C, effluent NO2-N concentrations can be higher than NO3-N
concentrations. There is little work on half-velocity coefficients for NO2-N oxidation by NOB. Manser et
al. (2005), in a comparative study, found that the NOB N substrate half-velocity coefficient was slightly
higher than that for AOB, but observations on activated sludge nitrification suggests a lower half-
velocity coefficient value. Based on the information in Table 4-3, the specific endogenous decay rate
may be assumed equal at 20°C for AOB and NOB
A possible set of kinetic coefficients for NOB are given in Table 4-6 and are selected to
represent the generally observed effects of temperature and DO concentration. The DO half-velocity
coefficient is much lower than that for AOB, as shown by Contreras et al. (2008) and Guisasola et al.
(2005). More research is needed on NOB kinetics, and site specific testing for model calibration may
give a different set of values than that shown in Table 4-6.
Table 4-6. Summary of NOB Nitrification Kinetic Coefficient Values
Parameter
Yield, Y
Specific endogenous decay
rate, b
Maximum specific growth
rate, umax
Half-velocity coefficient,
KNO
Half-velocity coefficient, K0
Units
g VSS/g N oxidized
g VSS/g VSS-day
g VSS/g VSS-day
mg/L
mg/L
20°C
Value
0.05
0.17
1.00
0.20
0.90
Temperature
correction value, 6
1.0
1.063
1.063
1.0
1.0
4.5.3 Effects of Temperature and Dissolved Oxygen on Nitrification Kinetics
The effects of temperature and aeration tank DO concentration on effluent NH4-N and NO2-N
concentrations are illustrated in Figures 4-2 and 4-3, respectively for a single completely-mixed activated
sludge (CMAS) aeration tank using Eqs. 4-9 and 4-10. No safety factors are built into these graphical
presentations. Figure 4-2 shows that a much longer SRT is needed at a lower temperature to achieve
low effluent NH4-N concentrations and that the effluent NO2-N concentration is always lower than the
NH4-N concentration. The example in Figure 4-3 shows that at the longer SRT needed to have an effluent
NH4-N concentration of 1.0 mg/L in a CMAS system at 10°C the effect of DO concentration on the
effluent NO2-N concentration is minimal, but at the higher 20°C temperature and lower SRT needed to
achieve an effluent NH4-N concentration of 1.0 mg/L, lower DO concentrations result in much higher
effluent NO2-N concentrations.
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15
20
5 10
SRT, days
Figure 4-2. Effect of SRT and temperature on effluent NH/-N and N02"-N concentrations using kinetic data in Table 4-6 and 4-4
for CMAS with no safety factor
NO2-N, 20
- NO2-N, 10
0.5 1.0 1.5
DO, mg/L
Figure 4-3. Effect of DO concentration on effluent N02"-N concentrations for SRTs at 10°C and 20°C that give an effluent NH4-N
concentration of 1.0 mg/L using kinetic data in Table 4-4 and 4-6 for CMAS with no safety factor
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It should be noted that the above examples on the effect of SRT, temperature and DO
concentration were based on steady state conditions of constant flow and influent concentrations.
Under typical diurnal flow and concentrations for domestic wastewaters, the SRT would have to be
increased to achieve the same effluent concentrations shown. Dynamic simulation models are used for
design to account for variable loading effects. (See Chapter 10 for detailed guidelines on using models
for design of biological nitrogen removal systems.) Typically the SRT may be 40 to 80 percent greater
than indicated by the figures.
The required SRT for nitrification can be reduced significantly for many applications by using a
series of aeration tanks (staged systems) versus a single aeration tank. The effect depends on the
influent nitrogen concentration to the aeration tank, temperature, and effluent NH4-N concentration
goal. The effect is less pronounced for systems with high return sludge recycle rates, such as MBRs, in
which the influent nitrogen is substantially diluted.
4.5.4 AOB and NOB Kinetics at High Temperature (SHARON® process)
The effect of temperature on the relative AOB and NOB kinetics is important in the SHARON®
(Single Reactor High Activity Ammonia Removal Over Nitrite) process used in sidestream treatment of
high ammonia anaerobic digestion centrate/filtrate from biosolids dewatering (Van Hulle et al. 2007;
Volcke et al. 2007). The process takes advantage of a high temperature condition (25°C to 35°C) in which
the specific growth rate of AOB is higher than that for NOB, so that at low enough operating SRTs the
NOB can be washed out and ammonia is oxidized to mainly nitrite (nitritation process). Low dissolved
oxygen conditions can also hinder NOB growth rates in the process. Based on the oxygen requirements
shown above of 3.22 g O2 used for NH4-N oxidation to NO2-N and 1.11 g O2 for oxidation of NO2-N to
NO3-N, the nitritation process reduces the energy required for complete nitrification by 26 percent. The
carbon needed for denitrification is also decreased.
Washout occurs when the SRT is lower than that needed to accommodate the maximum net
growth rate of the nitrifiers in the reactor. From Equations 4-6 and 4-8, the washout or minimal SRT is
as follows assuming excess N substrate and DO concentration:
SRTmin=^— Eq.4-18
The washout SRT for AOB and NOB as a function of temperature using the kinetic coefficients in
Tables 4-2 and 4-4, respectively, is compared to that claimed by Hellinga et al. (1998). Above 20°C, the
washout SRT for AOB is similar and below that for NOB.
A comprehensive model for the SHARON® process, including alkalinity effects, pH, free ammonia
and nitrous acid toxicity, and AOB and NOB kinetics was evaluated by Magri et al. (2007) with a bench
scale reactor operation at 35°C. The model high temperature u.max and decay coefficient (b) values for
AOB and NOB that best fit the lab results were 1.75 and 0.56 g/g-d and 0.23 and 0.04 g/g-d, respectively.
These values result in washout SRTs of 0.66 days for AOB and 1.92 days for NOB, which is in the range of
that shown for Figure 4-4 at 35°C.
Nutrient Control Design Manual 4-15 August 2010
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--•--AOB,H
-o--AOB,M
NOB,H
NOB,M
10.0 15.0 20.0 25.0 30.0 35.0
Temperature, °C
Figure 4-4. Effect of temperature on minimal washout SRT AOB, H and NOB, H from Hellinga et al. (1998) and AOB, M and NOB,
M from Tables 4-2 and 4-4.
4.6 Inhibitory Effects of Environmental Conditions on Nitrification
Effect ofpH
Ammonia oxidation rates decrease rapidly with decreasing pH below a pH of 6.8, and an optimal
pH is in the range of 7.5 to 8.0 (Tchobanoglous et al. 2003). The effect of lower pH on the ammonia
oxidation rate may be due to the reduction in free ammonia (NH3) concentration, as Suzuki et al. (1974)
reported that NH3-N is the true substrate for AOB. An operating pH of 7.0 to 7.2 is often used for design
and operation to assure reliable nitrification rates, and for some facilities it is necessary to add alkalinity
to maintain pH.
Effect of Free Ammonia and Nitrous Acid (HNO2)
For high ammonia strength wastewaters, such as from anaerobic digester centrate return,
animal feedlots, and industrial wastewater, the issue of NH3-N and HNO2 concentration inhibition on
AOB and NOB becomes important. Conditions that inhibit NOB activity to stop nitrification at NO2-N
have been modeled for sequencing batch reactors (Pambrun et al. 2006; Kyung and Choi 2001).
NH3-N is more inhibitory to NOB than to AOB (Peng and Zhu, 2006), while HNO2 is more
inhibitory to AOB. Anthonisen et al. (1976) and Turk and Mavinic (1986) found NOB inhibition at NH3-N
concentrations from 0.10 to 1.0 mg/L, whereas Mauret et al. (1996) found inhibition at 6.6 to 8.9 mg
NH3-N/L Wong-Chong and Loehr (1975) reported inhibition by NH3-N at 3.5 mg/L for unacclimated
Nutrient Control Design Manual
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August 2010
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bacteria and 40 mg/L after an acclimation phase. Kyung and Choi (2001) found 50 percent inhibition at
11.1 mg NH3-N/L
For AOB inhibition by NH3-N, Anthonisen et al. (1976) reported higher concentrations than for
NOB at 7.0 mg/L, and Abeling and Seyfried (1992) found complete loss of activity at a 20 mg NH3-N/L
concentration. However, with acclimation Wong-Chong and Loehr (1975) found stable AOB activity at
50 mg NH3-N/L. HNO2 concentrations found inhibitory to AOB were 0.065 to 0.83 mg/L by Anthonisen
etal. (1976).
The relative concentration of NH3-N and HNO2 are a function of pH and temperature and can be
calculated according to Anthonisen et al. (1976):
TAN(10pH)
NH,-N=—-7-i Eq.4-19
1/+1QPH
and Eq. 4-20
% =exp[6,334/(273 + T)]
N=
^n*10PHJ
and
Kn=exp[-2,300/(273+T)] Eq. 4-22
Where:
TAN = Total ammonia nitrogen, NH3-N + NH"4- N concentration, mg/L
HNO2-N = Free nitrous acid concentration as N, mg/L
NO2 - N = Nitrite concentration as N, mg/L
T = Temperature, °C
Ka = lonization constant for ammonium
Kn = lonization constant for nitrous acid
Values for Ka and Kn at 25°C are 10~9'24 and 10~3 4, respectively.
The above equations were used to calculate possible TAN and NO2-N concentrations that may
inhibit AOB and NOB at pH values from 6.0 to 8.0 at 20°C. Results are summarized in Table 4-7
Nutrient Control Design Manual 4-17 August 2010
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Table 4-7. NH4-N and NO2-N Concentrations that May Inhibit Nitrification as a Function of pH
at 20°C.
NH3 Inhibition
PH
6.0
6.5
7.0
7.5
8.0
AOB (7.0 mg NH3-N/L)
NH4-N, mg/L
17,100
5,400
1,700
550
178
NOB(0.10to3.5mgNH3-N/L)
NH4-N, mg/L
244-8,500
77-2,700
25-860
8-274
3-89
HNO2 Inhibition
PH
6.0
6.5
7.0
7.5
8.0
AOB (0.065 to 0.80 mg HN02/L)
N02-N, mg/L
25-311
80-986
253-3,118
801-9,861
2,533-31,185
Effect of salinity
Nitrifying bacteria are able to maintain activity under high saline concentrations. Campos et al.
(2002) found 100 percent nitrification at 13.7 g NaCI/L but a decline at close to 20 g/L
Inorganic and Organic Compound Inhibitors to Nitrification
Nitrifying bacteria are much more susceptible to inhibition than are heterotrophic bacteria.
They are affected by a wide range of organic and inorganic compounds that in many cases does not kill
them but greatly reduces their growth rate. Because of their sensitivity, they have been proposed as
indicators of the presence of toxic compounds at low concentrations (Blum and Speece 1991). Toxic
organic compounds include organic solvents, amines, phenolic compounds, alcohols, cyanates, ethers,
carbamates, and benzenes. The most significant heavy metals that are toxic to nitrifying bacteria at low
reactor soluble concentrations are nickel (0.25 mg/L), chromium (0.25 mg/L), and copper (0.10 mg/L)
(Tchobanoglous et al. 2003). Appendix B presents a list of known organic compounds that have been
identified as inhibitory to nitrification (WEF and ASCE 1998).
4.7 Denitrification Fundamentals
Denitrification is the biological reduction of nitrate or nitrite and can be assimilatory and/or
dissimilatory. Assimilatory denitrification involves the reduction of nitrate or nitrite to NH4-N for use in
biomass synthesis when NH4-N is not otherwise available. Most references to biological denitrification
for nitrogen removal refer to dissimilatory denitrification in which nitrate/nitrite is the ultimate electron
acceptor in the bacteria cell respiratory electron transport chain for the oxidation of various organic and
inorganic substrates.
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Nitrate reduction follows a series of intermediate products, nitrite (NO2~), nitric oxide (NO), and
nitrous oxide (N2O) to nitrogen gas (N2), with each step using a specific reductase enzyme in the
respiratory chain to transfer electrons. NO and N2O are nitrogen gases and the emission of N2O is of
great concern because it is one of the most significant greenhouse gases (See Chapter 2 for additional
discussion).
NO3->NO2->NO->N2O->N2 Eq. 4-23
Denitrification can be accomplished by heterotrophic bacteria oxidizing organic substrates,
heterotrophic nitrifying bacteria, and autotrophic bacteria. Heterotrophic bacteria are mainly
responsible for denitrification in the biological nitrification-denitrification processes. Most
heterotrophic bacteria responsible for biological denitrification use BOD in influent wastewater are
facultative aerobic bacteria with the ability to use elemental oxygen, nitrate, or nitrite as their terminal
electron acceptors for the oxidation of organic material. When oxygen is present, they will use oxygen
as the electron acceptor, but the reductase enzymes for denitrification are induced in the absence of
oxygen (Payne 1973). Autotrophic denitrification is of interest for higher temperature recycle streams
from anaerobic digestion centrate/filtrate with the ANAMMOX® bacteria process being the prime
example.
Microbiologists generally use the term "anaerobic" to describe biological reactions in the
absence of oxygen. To distinguish anaerobic conditions for which the biological activity occurs mainly
with nitrate or nitrite as the electron acceptor in biological nitrification-denitrification processes, the
term "anoxic" has been applied.
4.8 Microbiology of Denitrification
Heterotrophic bacteria capable of denitrification are very common in wastewater treatment and
include the following genera: Achromobacter, Acinetobacter, Agrobacterium, Alcaligenes, Arthrobacter,
Bacillus, Chromobacterium, Corynebacterium, Flavobacterium, Hypomicrobium, Moraxella, Nesseria,
Paracoccus, Propionibacteria, Pseudomonas, Rhizobium, Rhodopseudmonas, Spirillum, and Vibrio
(Tchobanoglous et al. 2003).
In many BNR process applications, a supplemental carbon source has been needed to (1)
provide sufficient carbon for nitrate/nitrite reduction for wastewaters with lower C/N ratios, (2)
accelerate denitrification rates to reduce tank volume requirements, or (3) provide a carbon source for
further nitrate/nitrite reduction in post-anoxic systems such as denitrification filters. Methanol (CH3OH)
has been commonly used because it is inexpensive, but because of its unique single-carbon compound
structure, it supports growth of a less diverse, more specific bacterial population. The methanol-utilizing
bacteria commonly found in denitrifying system are bacteria of the genus Hyphomicrobium
(Timmermans and Van Haute 1983; Sperl and Hoare 1971) and possibly Paracoccus denitrificans (Van
Verseveld and Stouthamer 1978). Recently, using stable isotope probing, Baytshtok et al. (2008)
identified a Methyloversatilis universalis strain in addition to Hyphomicrobium zavarzinili.
A heterotrophic nitrifying bacterium that can denitrify and has been studied often is Parococcus
pantotropha, which obtains energy by nitrate or nitrite reduction while oxidizing ammonia under
aerobic conditions. A readily available carbon source, such as acetate, is needed (Robertson and
Nutrient Control Design Manual 4-19 August 2010
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Kuenen, 1990). The conditions required for this form of denitrification are not practical in biological
wastewater treatment.
Denitrification has been observed for a number of autotrophic bacteria using nitrate/nitrite to
oxidize a variety of electron acceptors including zero valance iron and Fe(ll) by Paracoccusferrooxidans,
Paracoccus denitrificans, P. pantotrophus, and P. versutus (Kumaraswamy et al. 2006 Kielemoes et al.
2000), reduced sulfur compounds by Thiobacillus denitrificans (Bock et al. 1995) and ammonia by
Nitrosomonas eutropha, nitrosomonas europaea, and nitrosolobus multiformis (Poth and Focht 1985;
Zart and Bock 1998; Schmidt et al. 2003). Ammonia oxidation by nitrosomonas under anoxic conditions
has been shown to be slow and of little practical significance in activated sludge treatment (Littleton et
al.2003).
More recently Mulder et al. (1995) discovered a unique autotrophic bacteria in a denitrifing
fluidized bed reactor with the ability to oxidize ammonia to nitrogen gas using nitrite as the electron
acceptor. This biological reaction was termed ANAMMOX (ANaerobic AMMonia OXidation). The
autotrophic organisms capable nitrite reduction without carbon addition have been identified in the
ANAMMOX® process (Schmidt et al. 2003) as Candidatus Brocadia anammoxidans and Candidatus
Kuenenia stuttgartiensis and belong to the order Planctomycetales, a division within the domain
Bacteria (Strous et al. 1999). Under anaerobic conditions, ammonia is oxidized with the reduction of
nitrite to produce nitrogen gas. They are slow growing organisms, and the reaction is best accomplished
at temperatures above 25°C.
4.9 Metabolism and Stoichiometry of Heterotrophic Denitrification
Soluble organic substrates are consumed during heterotrophic denitrification, with a portion of
the substrate COD oxidized using nitrate or nitrite as the electron acceptor and the other COD portion
found in cell biomass. For denitrification of influent wastewater such as in pre-anoxic zones before
aeration, the soluble organic substrate consumed is from the following:
• The soluble degradable COD in the influent wastewater
• Hydrolysis of biodegradable particulate and colloidal COD in the influent wastewater
• Degradable COD release from endogenous decay
Oxidation reactions are shown as follows for wastewater represented as Ci0Hi9O3N (Tchobanoglous et
al. 2003) and common exogenous substrates.
Wastewater:
C10H1903N + 10N03^5N2 + 10C02 + 3H20 + NH3+100H- Eq. 4-24
Methanol:
5CH3OH + 6NO3^3N2 + 5CO2 + 7H2O + 6OH~ Eq. 4-25
Ethanol:
Eq. 4-26
Nutrient Control Design Manual 4-20 August 2010
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Acetate:
5CH3COOH + 8N03 -^4N2 +10CO2 +6H2O + 8OH" Eq. 4-27
In all these reactions, 1 mole of hydroxide alkalinity is produced per mole of NO3 reduced or one
equivalent OH per equivalent N. This equates to 50 mg alkalinity as CaCO3 per 14 mg N reduced or 3.57
mg alkalinity as CaCO3 produced per mg NO3-N reduced. The same alkalinity ratio also applies to NO2-N
reduction.
In the above reactions, nitrate serves as an electron acceptor for oxidation-reduction reactions
to oxidize the organic substrate as is done with oxygen as the electron acceptor. Because COD balances
are useful in biological process design to equate the biodegradable substrate consumed to oxygen or
electron acceptor needs plus biomass COD, it is convenient to understand the oxygen equivalent of NO3-
N and NO2-N. This can be done by comparing the half reactions for a mole of electron transfer as
follows:
Oxygen:
0.2502+H++e'^0.5H20 Eq. 4-28
Nitrate:
0.20NO3 + 1.2H+ +e" ->0.1N2 +0.6H2O Eq. 4-29
Nitrite:
0.33N02 +1.33H+ +e" -^0.17N2 +0.671^0 Eq. 4-30
For NO3-N, 0.20 moles are equivalent to the oxidation by 0.25 moles of oxygen, and thus the
oxygen equivalent of NO3-N equals (0.25*32)7(0.20*14) = 2.86 g O2 equivalent/g NO3-N reduced.
Similarly, the oxygen equivalent of NO2-N is 1.73 g O2/g NO2-N reduced. Thus, less substrate oxidation is
needed per unit of oxidized nitrogen removed for NO2-N reduction compared to NO3-N reduction so
that processes that stop nitrification at NO2-N need less carbon for denitrification.
4.10 Biological Denitrification Kinetics with Influent Wastewater
Denitrification rates in pre- or post-anoxic zones in BNR processes depend on many factors. A
semi-empirical approach has been to assess the rate in terms of a specific denitrification rate (SDNR) in
terms of g NO3-N reduced/g mixed liquor VSS (MLVSS)-day. Depending on the wastewater
characteristics, temperature, and design loading to a pre-anoxic zone, the SDNR may range from 0.03 to
0.20 g NO3-N/g MLVSS-day. For post-anoxic tanks it may range from 0.01 to 0.03 g NO3-N/g MLVSS-day,
where the denitrification rate is driven mainly by endogenous respiration rates (Tchobanoglous et al.
2003). The denitrification rate in an activated sludge reactor is affected by many factors and can be
determined using simulation models based on kinetic equations.
The organic substrate removal rate by denitrifiers determines the nitrate reduction rate. The
substrate removal rate is commonly described by Monod kinetics as follows and is affected by the
organic substrate, NO3-N and DO concentrations, and the biomass in the reactor that are facultative
nitrate reducers:
Nutrient Control Design Manual 4-21 August 2010
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Where:
rss = Substrate utilization rate in the reactor, mg/L-d
YH = Heterotrophic bacteria synthesis yield, g VSS/g COD used
Umax = Maximum specific growth rate of denitrifying heterotrophs, g VSS/g VSS-d
and is a function of temperature
Ss = Soluble degradable substrate concentration, mg COD/L
Ks = Substrate utilization half-velocity coefficient, mg COD/L
SNO = NO3-N concentration, mg/L
KMO = Nitrate half-velocity coefficient, mg/L
KO,H = DO inhibition constant for nitrate reduction, mg/L
S0 = DO concentration
H = Fraction of heterotrophic bacteria that can use nitrate in lieu of oxygen
XH = Heterotrophic bacteria concentration, mg/L
Additional equations and mass balances are required for the reactor to determine the reactor
soluble degradable COD concentration as a function of utilization with nitrate, soluble substrate
entering the reactor, and soluble substrate produced via hydrolysis of particulate and colloidal
degradable COD and from cell lysis due to endogenous decay. Wastewaters with a higher influent
soluble biodegradable COD fraction will have higher denitrification rates as the hydrolysis step to
convert particulate and colloidal COD to readily useable soluble COD is slower than the soluble COD
uptake rate.
The nitrate utilization rate is related to the fraction of biodegradable COD used that is oxidized
and incorporated into cell mass. A mass balance on degradable COD accounts for COD oxidized using
NO3-N and COD incorporated into cell synthesis as follows (Tchobanoglous et al. 2003):
1-1.42Y.
rNo= -rss Eq-4-32
Where:
rNO = NO3-N reduction rate, mg/L-day
and combining Eq. 4-31 and 4-32 gives the reactor NO3-N reduction rate as follows:
Eq. 4-33
The value for n is affected by the system operating conditions and how much of the growth
substrate is consumed under anoxic conditions. Stensel and Home (2000) showed that this can be at
least 0.80 for anoxic/aerobic BNR systems with significant BOD removal in a pre-anoxic zone. Literature
values for the heterotrophic kinetics coefficients in Eq. 4-31 vary widely and may be system specific for
the DO inhibition and half-velocity coefficients. Examples of coefficient values are given in Table 4-8.
Nutrient Control Design Manual 4-22 August 2010
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Table 4-8. Heterotrophic Bacteria Kinetic Coefficients in Anoxic/Aerobic Activated Sludge
Kinetic Model Parameter
Maximum specific growth rate, |amax
Temperature coefficient, 6
Heterotrophic bacteria synthesis yield, YH
Substrate utilization half-velocity coefficient, Ks
Nitrate half-velocity coefficient, KNO
DO inhibition constant, K0,H
Units
g VSS/g VSS-d
-
g VSS/g COD removed
mg COD/L
mg N03-N/L
mg DO/L
Typical Value
3.2
1.029
0.47
5.0
0.10
0.02
Source: Barker and Dold (1997)
The DO inhibition coefficient in Eq. 4-33 attempts to account for the inhibitory effect of oxygen
on denitrification. DO inhibition on denitrification has been shown at DO concentrations of 0.20 mg/L
by Skerman and Mori (1975) and Dawson and Murphy (1972). Oxygen inhibition is greater on nitrite
reduction than on nitrate reduction. There is less concern about pH effects on denitrification than for
autotrophic bacteria, though Dawson and Murphy (1972) showed a decrease in denitrification rates as
the pH was decreased from 7.0 to 6.0 in batch tests.
4.11 Denitrification Carbon Sources and Relative Consumption Ratios
Denitrifying bacteria need a readily available carbon food source, such as soluble BOD, to rapidly
convert nitrate to nitrogen gas. WWTPs that meet very low total nitrogen limits often use a secondary
anoxic zone in which supplemental carbon is added.
A general rule of thumb is that 4 g of wastewater influent BOD is needed per g ofNO3-N to be
removed through biological treatment (Tchobanoglous et al. 2003). When denitrification is needed
after nitrification, there is little BOD remaining so a supplemental carbon source is often needed.
Supplemental sources can be "internal," such as fermented wastewater or sludge, or "external" or
exogenous sources, such as purchased chemicals. The most common exogenous carbon source in use is
methanol; however, due to issues regarding its safety, kinetic rates, and availability, some wastewater
systems are using alternative carbon sources such as acetic acid, ethanol, sugar, glycerol, and
proprietary solutions depending on the needs of their particular facility (deBarbadillo et al. 2008).
Equation 4-32 can be rearranged to show the ratio of biodegradable COD required for complete
NO3-N reduction as a function of the biomass synthesis yield. This is similar to the consumptive ratio
advanced by McCarty et al. (1969), in which they showed that exogenous substrates with lower biomass
synthesis yields had lower consumptive ratios, and thus required less substrate addition relative to the
amount of NO3-N to be removed. The consumptive ratio for methanol was about 70 percent of that for
glucose, which had a higher biomass synthesis yield coefficient. A lower biomass synthesis yield means
that the bacteria oxidize a greater portion of the substrate to provide energy for biomass growth and
thus a greater portion of the COD used is oxidized by NO3-N with less ending up in sludge production.
CRN03 -
bCOD
2.86
N03-N 1-1.42YU
Eq. 4-34
Nutrient Control Design Manual
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August 2010
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Where:
CR|\io3 = S COD used/g NO3-N that is completely reduced in the anoxic zone.
Note that the value of 1.42(YH) is equal to the biomass yield in g biomass COD produced per g COD
consumed for an assumed biomass formula of C5H7N2O.
Based on the oxygen equivalent of the electron acceptor, a similar CR can be determined for
bacteria using NO2-N or oxygen in an anoxic zone. Oxygen may enter the anoxic zone in recycle flows or
in the influent wastewater. Thus the total exogenous chemical dose must account for other electron
acceptors entering the anoxic zone besides NO3-N.
bCOD 1.73
N02-N 1-1.42YH
Where:
CRi\io2 = g COD used/g NO2-N that is completely reduced in the anoxic zone.
Comparison of Equations 4-34 and 4-35 shows that the COD addition for NO2 reduction is about 60
percent of that for NO3 reduction.
bCOD 1.0
CR02= - = - Eq.4-36
DO 1-1.42Y.
n
Where:
CRo2 = g COD used/g DO that is completely reduced in the anoxic zone.
Figure 4-5 shows the effect of the biomass yield on the ratio of electron donor needed as COD to
NO3-N reduced to N2. Substrates that have lower biomass yields will require lower consumptive ratios.
Nutrient Control Design Manual 4-24 August 2010
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8.0
0.00 0.10 0.20 0.30 0.40 0.50
Biomass Yield, gVSS/gCOD
Figure 4-5. Ratio of COD required to N03-N completely reduce N03-N (CRN03) as a function of the biomass yield.
It should be noted that the biomass yield value in Equations 4-34 through 4-36 can also
represent the overall observed yield depending on the denitrification process design and where
endogenous decay occurs for the biomass grown on the electron donor. For example, if the exogenous
substrate is added in a tertiary application in which the reactor is mainly anoxic, the yield in the CR
equations can be taken as the observed yield (Yobs) for the system as follows:
CR
2.86
N03
Eq. 4-37
"obs ~~
1 + bSRT
Eq. 4-38
If the exogenous substrate is added to an anoxic reactor in a BNR system, Equation 4-34 is more
applicable for NO3 reduction as most of the endogenous decay for the methylotrophs will occur in other
aerobic and anoxic zones, which represent a large proportion of the total volume. However, the
estimated substrate addition should also take into account the nitrate demand from endogenous decay
by the mixed liquor in the anoxic zone, which would include that by methylotrophs and other
heterotrophs. The electron acceptor demand by the endogenous decay would consume a certain
amount of NO3-N. The consumptive ratio can then be applied to the remaining NO3-N to be removed in
that zone to estimate the exogenous substrate demand. A further consideration that will raise the
substrate demand is that not all of the exogenous substrate added will be used. Equation 4-33 shows
Nutrient Control Design Manual
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August 2010
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that there must be a finite substrate concentration to drive the nitrate reduction rate. For systems with
smaller anoxic tank volumes, a higher reactor substrate concentration is needed to remove the NO3-N at
a faster rate, which increases the loss of substrate to the reactor effluent and increases the total
exogenous substrate demand above that predicted by the consumptive ratio. Systems with higher
internal recycle flows through the anoxic zone receiving exogenous substrates will have a higher overall
feed COD/NO3-N ratio.
Values for biomass synthesis or observed yields are helpful for estimating exogenous organic
carbon feed quantities for nitrate/nitrite removal. Information on biomass yields for methanol (the
most commonly used exogenous substrate) and other possible exogenous substrates are summarized in
Table 4-9. In many cases, the yields are observed yields based on data fits from test data or calculated
using Eq. 4-37 and the reported COD/NO3-N ratio.
The higher the biomass yield, the higher will be the required dose for the exogenous substrate.
For methanol, the yields ranged from about 0.20 to 0.30 gVSS/g COD. Using Eq. 4-37 or Figure 4-4, the
CRN03 requirement for methanol is 4.0 to 5.0 g COD/g NO3-N. Methanol has a COD of 1.5 g COD/g
CH3OH, so the requirement based on methanol is 2.7 to 3.3 g CH3OH/g NO3-N. Field applications in the
range of 3.5 to 3.8 are common and account for the yield and the net effect of mixed liquor endogenous
decay, entering DO, and effluent methanol. Based on the yields in Table 4-9, the dose for acetate and
ethanol will be higher than that for methanol. CRN03 values ranged from 2.2 to 0.2 in an evaluation of 30
industrial waste organics by Monteith (1980).
Nutrient Control Design Manual 4-26 August 2010
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Table 4-9. Biomass Yields Reported for Exogenous Carbon Sources
Substrate
Methanol
Ethanol
Acetate
Glycerin
Glycerin
Glycerin
Glycerin
Glycerin
Glycerin
Glycerin
Glycerin
Glucose
Corn Syrup
Biomass Yield
g VSS/g COD
0.18
0.281
0.29
0.21
0.331
0.162
0.292
0.29
0.30
0.362
0.352
0.66
0.322
0.282
0.302
0.272
0.322
0.342
0.262
0.322
0.422
0.382
Reference
Stenseletal. (1973)
Dold etal. (2008)
Daiggeret al. 2007
Christensson etal. (1994)
Sobieszuk etal. (2006)
Carrera et al. (2003)
Cherchi etal. (2008)
Baytshtok etal. (2008)
Purtschert and Gujer (1999)
Christensson et al. (1994)
Cherchi etal. (2008)
Kujawa & Klapwijk (1999)
Bilyketal. (2009)
Bilyketal. (2009)
Bilyk et al. (2009)
Bilyk et al. (2009)
Bilyketal. (2009)
Bilyketal. (2009)
Bilyketal. (2009)
Bilyketal. (2009)
Bilyketal. (2009)
Bilyketal. (2009)
Observed yield
2From COD/N ratio
4.12 Denitrification Kinetics of Exogenous Carbon Sources
Because methanol has historically been the most commonly used exogenous carbon source, it
has received the most attention from industry experts regarding its degradation kinetics. More recently,
there has been greater interest in denitrification kinetics using ethanol, acetate, and glycerol.
4.12.1 Denitrification Kinetics with Methanol
Kinetic coefficients for methanol-utilizers have been evaluated based on a Monod model (Eq. 4-
10), and results for u.max, from various investigators are summarized in Table 4-10 along with
temperature effects according to Equation 4-17. The range of temperature correction coefficients
shown by Dold et al. (2008) are based on two different methods. The higher value is based on their high
food-to-microorganism (F/M) test procedure to obtain u.max. The authors believe that the endogenous
decay coefficient assumption may have biased the temperature correction value. The lower correction
value is based on observed specific denitrification rates for their methanol enrichments. In the study by
Baytshtok et al. (2008), u.max was also determined under conditions using nitrite as an electron acceptor
instead of nitrate, and a much lower specific growth rate was found at 0.28 g VSS/g VSS-day.
Although there are considerable results reported for the yield coefficient for methanol
utilization (Table 4-9), there is minimal information on the endogenous decay rate coefficient. Stensel et
al. (1973) reported a value of 0.04 g/g-d, and Purtchert and Gujer obtained a value of 0.25 g/g-d in their
Nutrient Control Design Manual
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August 2010
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model fit to denitrification kinetics with methanol. It should be noted that an endogenous decay
coefficient value (b) of 0.04 g/g-day was used by Nichols et al. (2007) and Dold et al. (2008). The use of
higher values in the data fitting methods would have resulted in higher u.max values.
The results reported by Purtschert and Gujer are of practical interest for BNR processes with
methanol feeding. Overdosing methanol to the anoxic zone or having a high methanol concentration in
the anoxic zone effluent to support high denitrification rates will then support growth of methylotrophs
under aerobic conditions. Their results suggest that the substrate utilization rates by methylotrophic
organisms grown under aerobic conditions may be lower than that grown under anoxic conditions.
Thus, kinetic models for methanol utilization in BNR processes may have to consider two methylotrophic
populations: those grown under (1) anoxic and (2) aerobic conditions.
Table 4-10. Reported Maximum Specific Growth Rates at 20°C and Temperature Coefficients
for Methanol Utilization under Anoxic and Aerobic conditions
Growth
Condition
Anoxic
Anoxic
Anoxic
Anoxic
Anoxic
Anoxic
Aerobic
Umax/
Anoxic
1.86
1.25
1.30
1.25
0.94
1.30
0.81
M-max/
Aerobic
1.72
3.88
Temp.
6
1.12
1.13
1.09-1.13
1.14
Reference
Stensel (1973)
Nichols etal. (2007)
Dold etal. (2008)
Cherchi etal. (2008)
Baytshtok et al. (2008)
Purtschert and Gujer (1999)
Purtschert and Gujer (1999)
Little information exists on the methanol half-velocity (Ks) coefficient value, and more
information is needed as this greatly determines the reactor methanol concentration when trying to
achieve sufficient denitrification rates while keeping the reactor effluent methanol concentration low.
Table 4-11 summarizes reported Ks values determined in laboratory kinetic studies. The results may be
highly dependent on data fitting methods, and further work is needed to evaluate denitrification at low
methanol concentrations to better define Ks.
Table 4-11. Reported Ks values for NO3-N reduction with methanol at 20°C
Growth
Condition
Anoxic
Anoxic
Anoxic
Anoxic
Ks,
mg COD/L
9.1
15.6
31.7
<1.0
Reference
Stensel (1973)
Cherchi et al. (2008)
Baytshtok etal. (2008)
Purtschert and Gujer (1999)
Because of the relatively slow denitrification kinetics for methanol and its rising cost, there is an
interest in using other exogenous substrates either long term, short term during cold weather
conditions, or in combination with methanol. In addition, where carbon is needed only periodically to
enhance denitrification rates, other carbon sources are more attractive as they do not require as long an
acclimation period compared to methanol.
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4.12.2 Alternative Exogenous Substrates and Denitrification Kinetics
Evaluation of alternative substrates has shown that denitrification rates are much faster with
acetate, ethanol, and glycerol than with methanol. Using the same high F/M test method as used for
methanol kinetics, Dold et al. (2008) obtained much higher values of u.max for acetate or sugar; at about
4.0 g/g-day at 20°C. Cherchi et al. (2008) using the same test technique as Dold et al. (2008) obtained a
20°C u.max value for acetate of 2.2 g/g-day. In comparing the maximum specific growth rate of bacteria
grown on methanol, acetate, and corn syrup (Table 4-12), Mokhayeri et al. (2006) found that acetate
and corn syrup resulted in rates that were about 2.5 times that for methanol grown bacteria at 13°C and
about 3.5 times faster at 19°C. Christensson et al. (1994) found a similar effect in comparing ethanol
and methanol. The maximum specific growth rate for ethanol compared to methanol was about 2.5
times faster at 15°C and about 2.3 at 25°C.
Table 4-12. Comparison of Maximum Specific Growth rates for Methanol, Acetate, and Corn
Syrup at High and Low Temperatures
Reference
Mokhayeri et al.
(2006)
Christensson et al.,
1994
Substrate
Methanol
Acetate
Corn Syrup
Methanol
Ethanol
umax, g VSS/g VSS-day
13°C
0.5
1.2
1.3
15°C
0.8
1.9
19°C
1.0
3.7
3.5
25°C
2.1
4.8
Mokhayeri et al. (2008) also found a similar rate comparison using three enrichments grown on
methanol, acetate, and ethanol by comparing SDNRs with unlimited substrate at 13°C. The ratio of the
acetate to methanol SDNRs was 3.4 and for ethanol to methanol it was 3.3. Fillos et al. (2007) presented
the equations shown below to describe SDNRs with methanol or ethanol addition to an activated sludge
BNR process as a function of temperature. They showed appreciable differences in SDNR with ethanol
versus methanol, but in contrast to the above, the relative denitrification rates with ethanol versus
methanol was not as large and greater at high temperature; ethanohmethanol SDNRs are 2.2 at 20°C
and 1.8 at 10°C. After a 50-day acclimation period in a full scale and pilot scale anoxic-aerobic BNR
systems, Hallin and Pell (1998) found that the system with ethanol addition had an SDNR that was about
2.2 times the system with methanol addition.
Methanol SDNR (Fillos et al, 2007):
SDNRT=0.0738(1.11)(T"20)
Ethanol SDNR (Fillos et al., 2007):
SDNRT= 0.161(1.13)(T"20)
Eq. 4-39
Eq. 4-40
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4.12.3 Acclimation Time and Degradative Ability of Denitrifying Bacteria with Exogenous
Substrates
For some BNR applications, the need for an exogenous substrate may be seasonal to increase
denitrification rates at low temperatures or intermittent due to wet weather conditions or low weekend
loads. In such cases, the ability of the BNR activated sludge mixed liquor to rapidly respond to the
exogenous substrate addition is desired. For methanol additions, a considerable acclimation period is
needed to fully utilize the methanol added and to achieve the full denitrification rate possible with the
methanol addition. Acclimation time is not a significant issue with some other exogenous substrates,
such as ethanol and acetate. In full-scale and pilot-plant BNR system studies treating municipal
wastewater, Hallin and Pell (1998) found an immediate response with ethanol and acetate addition. In
contrast, a time period of over 50 days was required to reach full degradation rates with methanol as
shown by batch tests using the BNR mixed liquor with excess substrate. Table 4-13 shows the increased
denitrification rates over time relative to the BNR activated sludge without substrate addition.
Methanol addition did have some mild immediate effect to the denitrification rate potential, although
approximately 1 SRT (26-day SRT for system) was needed to fully develop the methanol-degrading
population. Nyberg et al. (1992, 1996) also showed the need for significant acclimation time of two- and
three-SRTs (50 and 70 days) with methanol feeding before the methylotrophic population reached its
full degradative capacity.
Table 4-13. For BNR Activated Sludge, Ratio of Denitrification Rate with Substrate Addition to
Denitrification Rate with No Addition
Time Period
Dayl
Day 50
Substrate
Methanol
1.2
2.0
Ethanol
4.0
4.1
Source: Hallin and Pell (1998)
Another important observation by Hallin and Pell (1998) was that the bacteria populations
developed by ethanol and methanol feeding could also degrade other exogenous substrates (Table 4-
14). The ability of the methanol-fed BNR activated sludge to degrade ethanol was also observed by Dold
et al. (2008) in which they commented that it violated the rule that methylotrophs can only use single
carbon compounds. However, Sperl and Hoare (1971) noted that the methylotroph Hyphomicrobium
sp. can denitrify with ethanol, propanol, butanol, acetate, formate, formaldehyde, methylamine, and
glycerol. It should be noted also that Dold et al. (2008) obtained a u.max value for ethanol of 0.37 g/g-d at
9.8°C, which is close to that predicted for methanol at the same temperature; however, an ethanol
enriched activated sludge should have a much higher rate. Nyberg et al. (1996) raised the possibility of
using ethanol with methanol to accelerate the acclimation time of a methanol-fed denitrification
system. Cherchi et al. (2009) found that a methanol acclimated biomass could also use acetate,
ethanol, and MicroC™ without an acclimation period, although glucose did require an acclimation
period. MicroC™ acclimated biomass was found to be able to denitrify using methanol, ethanol,
acetate, and glucose without an acclimation period. They reported that acetate acclimated biomass
could not use any other carbon sources to denitrify without an acclimation period.
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Table 4-14. Ratio of Denitrification Rates for Other Substrates at Day 50 with Ethanol or
Methanol Addition Versus no Addition
Test Substrate
Methanol
Ethanol
Acetate
Propionate
Butyrate
Glucose
Ethanol Fed Mixed Liquor
7.0
4.1
2.4
1.9
2.3
1.8
Methanol Fed Mixed Liquor
2.0
1.7
1.0
0.8
0.8
0.8
Source: Hallin and Pell (1998)
For the results shown above in Table 4-14, the ethanol utilization rate by the methanol-fed
system was 85 percent of its methanol utilization rate. Cherchi et al. (2008) observed an ethanol
degradation rate that was 18 percent higher than the methanol utilization rate for a methanol fed
enrichment. The methanol-utilizing bacteria contain methanol dehydrogenase enzymes that can act on
a number of alcohols.
4.13 Specific Denitrification Rates (SDNR)
SDNR values have been commonly used to estimate the size of pre- and post-anoxic zones for
nitrogen removal. They provide a simple first cut approximation of the anoxic reactor sizing. More
precise calculations for anoxic reactor size and the effect of stages are commonly done today with
commercial simulation model software that models the fate of degradable particulate and soluble
substrates, active biomass, and nitrogen species and the effect of temperature and DO concentration.
See Chapter 10 for guidance on using simulation models for the design of biological nutrient removal
systems. This section presents the fundamental equations used in models to size reactors and presents
typical values for kinetic parameters.
The required anoxic volume for denitrification can be estimated from SDNRs with the following
equation.
SDNR(X)
Where:
V = Anoxic zone volume, m3
SDNR = Specific denitrification rate, mg NO3-N removed/mg MLVSS-day
X = MLVSS concentration
NOR = Nitrate removal rate in anoxic zone equal to nitrate fed rate to anoxic zone via internal and
return sludge flows or from upstream nitrification zone.
Typical values for SDNR are summarized in Table 4-15. Temperature corrections for methanol addition
have been given previously. For wastewater feed only, reported values range from 1.03 to 1.08
(Dawson and Murphy, 1972, Ekama et al., 1984).
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Table 4-15. Range of reported SDNR values in BNR activated sludge treatment
Condition
Pre-anoxic
Post-anoxic
With methanol added
SDNR, 20°C
g NOs-N/g MLVSS-d
0.05-0.15
001-0.04
0.10-0.25
Source: Tchobanoglous et al. 2003
The following equation was used to fit SDNR data versus the pre-anoxic zone food to mass ratio
(Ib BOD applied/lb MLVSS-day) for a Bardenpho process at 18°C with no primary treatment (Burdick et
al., 1982).
SDNR=0.03(F/M) +0.029
Eq. 4-42
QSo
M
Eq. 4-43
Where:
SDNR20 = Specific denitrification rate at 20°C, g NO3-N/g MLVSS-d
F/M = BOD food to mass ratio to the anoxic zone, g BOD/g MLVSS-d
Q = Influent flowrate, m3/day
So = Influent BOD concentration, mg/L
X = MLVSS concentration of the anoxic zone, mg/L
VA = Anoxic zone volume, m3
Equation 4-42 has to be adjusted for other conditions of SRT and wastewater characteristics that will
result in a different active biomass fraction in the MLVSS. The active biomass fraction for the data fit was
estimated at 0.30 using Equation 4-44. Thus, Equation 4-42 is adjusted as follows:
Eq. 4-44
Eq. 4-45
Where:
Fb
YH
bT
= Active biomass fraction of MLVSS
= Heterotrophic biomass synthesis yield, 0.47 g VSS/g BOD removed
= Endogenous decay rate at MLVSS temperature, g VSS/g VSS-d
= Endogenous decay rate at 20°C, 0.10 g VSS/g VSS-d
= Influent inert VSS fraction, g VSS inert/g BOD
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The influent inert solids concentration greatly affects the active biomass fraction of the MLVSS.
Values for Y, are site specific but generally range from 0.10 to 0.30 for secondary influent wastewaters
with primary treatment and 0.30 to 0.50 without primary treatment (Tchobanoglous et al., 2003).
Temperature corrections applied to the SDNR in Equation 4-44 and for the endogenous decay
rate (bT) are as follows:
bT=b20(1.029(T-20)) Eq.4-46
SD/V/?T = SDNR20(1.Q~/J~20)) Eq. 4-47
For post-anoxic systems or for the anoxic zone of oxidation ditch processes, Refling and Stensel (1978)
used the following equation to estimate the SDNR.
SDNRET= n(AN) f—I Eq.4-48
EJ 2.86(YNRSRTj
Where:
SDNRE/T = SDNR in anoxic zones following nitrification and with no exogenous carbon addition, g NO3-
N/g MLVSS-d
H = Fraction of heterotrophs capable of nitrate reduction, 0.50 to 0.85
AN = Net oxygen requirement by heterotrophs, g O2/g BOD removed
YN = Net heterotrophic biomass yield, g VSS/g BOD removed
YN and AN are calculated with Equations 4-49 and 4-50, respectively.
Y = u Eq. 4-49
N l + bTSRT
AN=1.6-1.42(YN) Eq.4-50
4.14 Simultaneous Nitrification-Denitrification
Simultaneous nitrification-denitrification (SNdN) refers to a condition in activated sludge (or
biofilm) processes in which the positive bulk liquid DO concentration is low enough that the DO diffusing
into the floe is removed before it can penetrate the entire floe depth. Thus, nitrification is occurring on
the exterior portions of the floe and denitrification is occurring in the anoxic, interior portion.
SNdN commonly occurs in oxidation ditches with sufficient SRT and with aerobic and anoxic
zones as flow moves downstream from aeration in the ditch channels. Nitrite accumulation does not
typically occur under low DO conditions when SRTs are maintained at appropriate values, meaning that
the AOB and NOB populations remain balanced. Park et al. (2002) evaluated the AOB population in an
aerated-anoxic Orbal process and found that Nitrosospira-like organisms were a major contributor to
ammonia oxidation. Nitrosospira have a high affinity for oxygen which allows them to grow under low
DO conditions.
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Daigger et al. (2007) illustrated that the floe size for a given system is an important parameter
that can affect the DO penetration and amount of denitrification. The DO concentration that is possible
for SNdN depends on a number of factors including the mixed liquor concentration, temperature,
oxygen uptake rate, substrate loading, and floe size. In SNdN systems, the nitrification and
denitrification rates are lower than that for systems with separate nitrification zones with higher DO
concentration and separate denitrification zones with no DO present. Thus, the reactor volumes for a
SNdN system should be larger than that for a system with a separate anoxic denitrification zone and
higher DO aerobic nitrification zone.
4.15 Metabolism, Stoichiometry and Kinetics of ANAMMOX®
The autotrophic bacteria anaerobic oxidation of ammonia with nitrite has been termed the
ANAMMOX® process or Anammox reaction (Strous etal., 1998). The Stoichiometry for the biological
reaction given by Strous et al. (1998) is as follows, showing that 1.32 moles of NO2-N are reduced per
mole of NH4-N oxidized:
NH+ +1.32NOI1 +0.13H+ +0.066HCO:1 -»
Eq.4-51
1.02N2 +0.26N031 +0.066CH2005N015 +2.03H20
These are slow-growing organisms with doubling times of about 11 days at 30°C (Jetten et al., 1999).
Biokinetic parameters with selected values have been assembled by Capuno et al. (2008) as part
of a model development for the process in a biofilm reactor (Table 4-16).
Table 4-16. ANAMMOX® Bacteria Biokinetic Parameters at 30 C
Parameter
Maximum specific growth rate
Growth yield
NH4-N half-velocity coefficient
NO2-N half-velocity coefficient
Oxygen inhibition coefficient
Specific Endogenous decay rate
Units
g VSS/g VSS-d
g VSS/g N oxidized
mg/L
mg/L
mg/L
g VSS/g VSS-d
Value
0.08
0.11
0.07
0.05
0.01
0.003
Source: (Capuno et al., 2008)
The process is sensitive to nitrite concentration. An NO2-N concentration of 4.8 mg/L was found
to decrease anammox activity by Wett et al. (2007). Musabyimana et al. (2008) reported that a
preferred pH operating range is 7.0 to 7.7 and that activity could be sustained at 50 mg/L NO2-N, though
inhibited. They also noted that nitrite inhibition was reversible.
The original Anammox process relied on a step process in which nitrite is produced under
aerobic conditions in a portion of the ammonia rich stream and then the streams are combined and
subject to anaerobic conditions. Other process modifications, such as the DEMON, CANON, and OLAND
processes are described in Chapter 6.
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4.16 Impacts on Sludge Production and Handling
Less sludge production and better settling and thickening sludge is found for activated sludge
systems using anoxic/aerobic treatment for nitrogen removal versus aerobic treatment only. The sludge
production is less because the biomass yield from BOD consumption using nitrate as an electron
acceptor instead of oxygen is about 40 percent lower. The net reduction in sludge production is site
specific and depends on the portion of the influent BOD that is oxidized via biological denitrification. For
many applications, reductions of biomass production by 10 to 20 percent are possible. The impact this
will have on total sludge production by a treatment plant will depend upon how much waste sludge is
produced by other treatment units such as primary clarifiers and chemical treatment with precipitating
chemicals.
Implementation of biological nitrogen removal at conventional activated sludge plants improves
the sludge thickening characteristics due to its ability to decrease the amounts of filamentous bacteria in
the activated sludge. The effect of additional solid production from adding an exogenous carbon source
to increase denitrificaton rates must be considered as part of the overall sludge production. Solids
produced from nitrogen removal processes generally thicken and dewater well and show no negative
impact on any solids processing system.
4.17 Effluent Dissolved Organic Nitrogen
As effluent limitations have become more stringent, the DON fraction in the plant effluent has
become increasingly important. Most BNR processes can remove 80 to 95 percent of the inorganic
forms of nitrogen (ammonia-nitrogen, nitrate, and nitrite), but are less efficient at removing the residual
organic fractions (Sedlak 2007). The concentration of DON in the plant effluent (also called effluent
DON or EDON) depends on influent concentration and specific treatment process and varies from plant
to plant. Based on effluents from 188 BNR plants, Pagilla (2007) reported concentrations typically
between 0.5 and 1.5 mg/L with some values as high as 2.5 mg/L Likewise, Bratby et al. (2008) found
EDON concentrations between 0.4 and 2.2 mg/L with an average value of 1.8 mg/L based on a
compilation of the available literature. The EDON fraction is not as important for plants trying to
achieve a TN effluent limit of 10 mg/L, but becomes a much larger percentage of effluent TN and more
problematic for plants attempting to meet a 5 mg/L or 3 mg/L limit.
The text box below provides information on EDON in the form of questions and answers. One
very important and timely question for the industry is whether EDON can degrade in the natural
environment to bioavailable inorganic nitrogen species. Bioavailable forms can be used by algae and
other phytoplankton and thus, contribute to the problem of eutrophication. Although much of EDON
has been thought to be recalcitrant, meaning that it is inert and not bioavailable in the natural
environment, Mullholland et al. (2007) noted that photochemical reactions and salinity conditions can
convert recalcitrant material to inorganic forms of nitrogen such as nitrite and ammonium. Bacteria can
also play an important role in the breakdown of EDON (Sedlak and Pehlivanoglu 2007). Summarizing
results of several research projects, the Water Environment Research Foundation (WERF) (2008)
estimates that for those WWTPs discharging into freshwater watersheds, 20 to 60 percent of EDON
could be converted to bioavailable forms. The authors note, however, that the proportions are very site
specific and additional research is needed in this area.
Nutrient Control Design Manual 4-35 August 2010
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Questions and Answers Regarding Effluent Dissolved Organic Nitrogen (EDON)
What are the sources of EDON? There are two primary sources:
(1) Influent DON that is recalcitrant and passes through the plant unchanged or influent
biodegradable DON that is not removed during treatment. DON in domestic wastewater can
be present as amino acids, proteins, aliphatic N compounds, synthetic compounds such as
EOT A, and humic organic substances (WEF 2008).
(2) DON can also be produced during biological treatment processes within the plant through
cell metabolic processes, cell decay, and cell lysis.
What are its characteristics? EDON is composed primarily of degraded amino sugars, peptides, and
porphyrins (Leenheer et al. 2007 as cited in Bratby et al. 2008). The molecular weight varies from
source to source but has been reported to be relatively small overall (Bratby et al., 2008). Sedlak and
Pehlivanoglu (2007) hypothesized that the high molecular weight portion (> 1,000 Daltons) is not
biologically available, whereas a large portion of the low molecular weight portion (< 1,000 Daltons)
may be biodegradable.
How is it measured? EDON is usually determined as the difference between the measured effluent
soluble Kjeldahl nitrogen and effluent ammonia with filtration of both samples through a 0.45 micron
filter. Because of challenges associated with measuring fractions in effluents with very low TN
concentrations, researchers are investigating other methods. Sattayetewa and Pagilla (2008)
reported success using a persulfate disgestion method and second derivative ultraviolet
spectrophotometric (SOUS) combined method. Currently, there is no standard method for
differentiating between the inert portion and the portion that can be converted to bioavailable
inorganic nitrogen in the watershed.
How is it removed at the plant? DON can be removed through hydrolysis and ammonification.
Removal is dependent on temperature and solids residence time.
What are its environmental concerns? The main concern in the wastewater industry is the
bioavailability of effluent DON when it is a large fraction of the effluent total nitrogen concentration.
For more information, see the WEF Nutrient Compendium, available online at
http://www.werf.org/AM/Template.cfm?Section=Content Folders&CONTENTID=8726&TEMPLATE=/
CM/ContentDisplay.cfm
Nutrient Control Design Manual 4-36 August 2010
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Chandran, K., Z. Hu and B. F. Smets. 2008. A Critical Comparison of Extant Batch Respirometric and
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Experimental Measurement, Modeling, and Implications for Simultaneous Nitrification and
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5. Principles of Biological Phosphorus Removal
Chapter 5 covers:
5.1
5.2
5.3
5.4
5.5
5.6
5.7
5.8
Overview of the Biological Phosphorus Removal Process
Substrate Requirements
Sources of Volatile Fatty Acids
Environmental Conditions
Kinetics
Important Design and Operational Considerations
Impacts on Sludge Processing and Handling
References
5.1 Overview of the Biological Phosphorus Removal Process
Biological phosphorus removal (BPR) in wastewater treatment is accomplished by encouraging
the growth of phosphate accumulating organisms (PAOs). PAOs are heterotrophic bacteria that occur
naturally in the environment and in aerobic activated sludge. The growth of PAOs is encouraged by
cycling them between anaerobic and aerobic conditions.
In the presence of oxygen (i.e., aerobic conditions), PAOs obtain energy from stored food and
uptake large amounts of phosphorus into their cells, which they store as polyphosphates. These
polyphosphates contain high-energy bonds and function like energy storage batteries. In the absence of
oxygen (i.e., anaerobic conditions), PAOs can break the polyphosphate bonds and use the resulting
energy to uptake easily biodegradable compounds, namely short chain volatile fatty acids (VFAs)1. PAOs
polymerize and store the VFAs in their cells as intermediate products known as poly-B-hydroxy-
alkanoates (PHAs), of which the most common is poly-B-hydroxy-butyrate (PHB). When oxygen
becomes available again (i.e., aerobic conditions), they can metabolize the PHAs to generate energy and
uptake phosphorus (in the form of phosphate) and store the excess amount. See Figure 5.1 for a
conceptual representation of the theory of BPR.
Aerobic
Figure 5-1. Theory of BPR in activated sludge.
1 For biological nutrient removal (BNR), it is important to distinguish between anaerobic conditions in where no
oxygen is present and anoxic conditions where oxygen is available in combined form only (e.g., NO2", NO3") and
there is no free oxygen. Anoxic conditions are sufficient for denitrification, whereas anaerobic conditions are
required for BPR.
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What makes PAOs special? PAOs have a competitive advantage over most aerobic bacteria
because they can uptake and store food (i.e., VFAs) under anaerobic conditions. Most other
microorganisms must wait until oxygen is available to uptake VFAs. One notable exception is glycogen
accumulating organisms (GAOs), which can also uptake VFAs in an anaerobic environment using energy
stored in glycogen. GAOs compete with PAOs in the anaerobic zone; however, PAOs nearly always out-
compete GAOs in temperate and cold-zone treatment plants. There is still a debate among researchers
about the conditions likely to favor GAOs over PAOs. Summarizing a number of publications, it would
appear that the following conditions favor the growth of GAOs over that of PAOs:
• Temperature over 28° C.
• Stronger waste with low nitrogen content.
• Polysaccharides such as glucose are the dominant feed to the anaerobic zone.
• Low pH in the aerobic zone.
Other factors that may favor GAOs but need further confirmation are high solids retention time (SRT),
longer non-aerated zones, and periods of intermittent low organic substrate levels in the plant influent.
Note that some GAOs are always present in the anaerobic zones of municipal BPR plants. Typically their
impacts are not noticeable because the PAOs still obtain enough VFAs to remove the phosphorus to the
desired effluent concentration.
The key steps for achieving BPR at wastewater treatment plants (WWTPs) are illustrated in the
simplified diagram in Figure 5-2 and summarized below.
1) Return activated sludge from the secondary clarifier, which contains PAOs, is added to the
influent wastewater.
2) In the anaerobic zone, PAOs break polyphosphate bonds to generate energy. They use this
energy to take up VFAs and store them as PHA compounds such as PHB. When PAOs break the
polyphosphate bonds, they release phosphate ions (PO4~3) in their cells. Because each
phosphate molecule is negatively charged, it must first bond with positively charged ions such as
magnesium or potassium to move across the cell membrane and be released into the water.
3) When the mixed liquor (i.e., wastewater and return sludge) enters the aerobic zone, the PAOs
use oxygen to metabolize the stored PHAs to generate energy for growth and maintenance.
They store the excess energy by taking up phosphate ions along with magnesium, potassium,
and other positive ions into their cells and forming polyphosphates. This "luxury uptake" of
phosphorus results in more being removed in the aerobic zone than was released in the
anaerobic zone.
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4) As the water enters the clarifier, PAOs settle to the bottom with the rest of the activated sludge.
The phosphorus stored in the PAOs is removed with the waste activated sludge, thereby
resulting in a net removal of phosphorus during treatment2.
Influent-
©
AnerobicZone
PAOs store
volatile fatty acids
as PHAs.
Aerobic Zone
PAOs break down PHAs for energy and to
obtain carbon for growth. They use this
energy for phosphorus uptake and cell
growth
Effluent
Return Activated Sludge
Phosphorus Wasted with
Waste Activated Sludge
Figure 5-2. BPRataWWTP
5.2 Substrate Requirements
The availability of readily biodegradable organic carbon (i.e., VFAs) in the anaerobic zone is
critical to the success of BPR. Table 5-1 shows the VFAs typically found in fermented wastewater and
the observed ratio of mg/L phosphorus uptake per mg/L of each VFA consumed during the treatment of
municipal wastewater (Abu-Ghararah and Randall 1991). Acetic acid and propionic acid are the
dominant VFAs in domestic wastewater, with the other forms present in minimal concentrations.
Common percentages from the fermentation of municipal wastewater settled solids are 60 percent
acetic, 30 percent propionic, and 10 percent of the remaining four. Large variations occur, however.
Propionic acid, for example, may vary from less than 20 to more than 50 percent, and isovaleric acid
may be as high as 10 percent. Research has shown that GAOs do not grow well on propionic relative to
PAOs (Chen and Randall 2004; Lopez-Vazquez et al. 2009). Therefore, the ratio of propionic to acetic
affects the performance of BPR, with higher amounts of propionic relative to the acetic producing
improvements.
If wasting is directly from a reactor rather than from the clarifier as is practiced at some activated sludge plants
and at plants with membrane bioreactors, the activated sludge should be wasted from an aerobic zone, preferably
when the sludge phosphorus concentration is at a maximum (i.e. before significant release occurs from extended
aeration).
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Table 5-1. Volatile Fatty Acids Typically Found in Fermented Wastewater
Volatile Fatty Acid
(VFA)
Acetic acid
Propionic acid
Butyric acid
Isobutyric acid
Valeric acid
Isovaleric acid
Chemical Formula
CH3-COOH
CH3-CH3-COOH
CH3-CH3-CH3-COOH
CH3-CH3-COOH-CH3
CH3-CH3-CH3-CH3-COOH
CH3-CH3-COOH-CH3-CH3
Phosphorus Uptake / VFA as
COD Consumed
0.37
0.10
0.12
0.14
0.15
0.24
Note: Acetic acid and propionic acid are the most common VFAs found in fermented wastewater.
Source: Abu-Ghararah & Randall (1991) as presented in Table 4.2 from WEF and ASCE (2006)
Table 5-2 summarizes the minimum ratios of substrate to total phosphorus to obtain effluent
phosphorus concentrations of less than 1 mg/L that have been reported in the literature. Evaluating the
chemical oxygen demand to total phosphorus (COD:TP) and five-day biochemical oxygen demand to
total phosphorus (BOD5:TP)3 ratios will provide only a rough approximation of BPR capabilities, but will
provide a useful rule of thumb for estimations. Evaluating the ratio of readily biodegradable COD to TP
(rbCOD:TP ratio) of the process influent is a more reliable way of assessing the BPR capabilities of a
wastewater plant and of determining if additional substrate is required to achieve the desired effluent
phosphate concentration. Analysis of the rbCOD:TP ratio is superior to analysis of VFA:TP because much
of the non-VFA rbCOD can be fermented to VFA in the anaerobic zone as long as there is sufficient SRT
(typically at least 1 to 2 days but varies with temperature; WEF and ASCE 2006). Hence, designers may
under-predict performance and over-predict the need for additional substrate if only VFA:TP ratios are
considered. See the text box following Table 5-2 for information on how to estimate the rbCOD:TP ratio.
Table 5-2. Minimum Ratios for Achieving Total Phosphorus Effluent Concentration of less than 1.0
mg/L
Substrate Type to Total
Phosphorus
COD:TP
BOD:TP
rbCOD^P1
VFA: TP
Recommended Minimum Ratio
40-45
20
10-16
4-16
References
WEF and ASCE 2006
WEF and ASCE 2006
Barnard et al. 2006
Neethlingetal. 2005
Most accurate
COD = chemical oxygen demand; BOD = biochemical oxygen demand; rbCOD = readily biodegradable chemical oxygen demand;
VFA = volatile fatty acid; TP = total phosphorus
Unless otherwise indicated, "BOD" is based on the 5-day test. Ultimate BOD (BODU) is based on the 20-day test
per standard methods (APHA, AWWA, and WEF 2005).
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How can I determine the rbCOD:TP ratio at my wastewater plant?
Total Phosphorus (TP). Information on influent TP levels should be available for your treatment
plant. TP in domestic wastewater typically ranges between 4 and 8 mg/L but can be higher
depending on industrial sources, water conservation, and whether or not a phosphate detergent ban
is in place. It is typically measured using colorimetric analysis or ion chromatography and requires a
digestion step to convert polyphosphate and organic phosphorus to orthophosphate. TP loading
typically follows a diurnal pattern with highest concentrations during the day. Influent
concentrations are usually lower during peak wet weather flow events.
Readily Biodegradable Chemical Oxygen Demand (rbCOD). The literature describes several
experimental methods that can be used to estimate rbCOD. After extensive review, Melcer et al.
(2003) concluded that the Aerobic Batch Method is the most practical of the bioassay methods, has
the least requirement for analytical laboratory facilities, and lends itself to onsite use at WWTPs. For
detailed instructions on how to perform this method, refer to Melcer et al. A second, less time-
consuming method was published by Mamais et al. (1993) and is based on the assumption that truly
soluble degradable COD is representative of the rbCOD. The method requires removal of solids and
colloids by flocculation followed by filtration through a 0.45 micron filter prior to COD analysis. The
difference in results for an influent and effluent sample is taken to be rbCOD. Melcer et al. found
that although the Mamais method is comparable to bioassay methods, results can be up to 5 percent
higher. Influent rbCOD can vary significantly between seasons in temperate climates with ranges
from 25 to 125 mg/L. See Chapter 10 for additional information on estimating COD fractions.
Research results suggest that the instantaneous COD:TP ratio is more important than the overall
average (Neethling et al. 2005). Short-term drops in the BOD:TP ratio in the primary effluent to below
that required for the desired quantity of phosphorus removal correlated well with rises in effluent
phosphorus. Intermittent recycles of phosphorus-rich return streams may cause short-term variability
in the BOD:TP ratio. Weekend changes in the BOD:TP ratio also can affect performance.
5.3 Sources of Volatile Fatty Acids
Section 5.2 described how fermentation of the non-VFA portion of rbCOD can occur in the
anaerobic zone of WWTPs to produce additional VFAs. Other sources of VFAs include:
• Fermentation in the wastewater collection system
• Fermentation at the treatment plant (requires new process equipment)
• Commercial sources
Each source is discussed separately below.
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5.3.1 Fermentation in the Collection System
Fermentation of wastewater under anaerobic conditions occurs to at least some extent in all
wastewater collection systems. Factors that favor anaerobic conditions are:
• Long detention times (e.g., in relatively flat collection mains or pump station wet wells)
• High strength waste
• Warm temperatures
Force mains are also excellent fermenters for the production of VFAs because they flow full and there is
no air-water interface for oxygen transfer. Often, the concentration of influent VFAs varies from month
to month depending on temperature and flow conditions in the collection system. Further changes will
occur within the plant depending on the extent and type of pretreatment, the types of sludge processing
equipment, and the utilization of anaerobic digestion.
In addition to VFAs, fermentation produces sulfide. Sulfide is converted to hydrogen sulfide
(H2S) when it enters air such as the head space in gravity sewers or in force main discharge manholes.
Hydrogen sulfide has an objectionable odor, and its release can cause corrosion problems when it is
converted to sulfuric acid by bacteria on the pipe wall.
Some techniques used for odor and corrosion control in wastewater collection systems do so by
reducing anaerobic conditions, which can have detrimental impacts on BPR. Kobylinski et al. (2008)
reported that chemical oxidation, nitrate addition, pH control to greater than 11, super oxygenation,
and vapor-phase odor control can slow or stop the production of VFAs. Other techniques such as iron
addition and pH control to between 9 and 10 do not adversely impact the production of VFAs. Table 5-
3 provides a summary of the effect of odor and corrosion control techniques on VFA formation in
wastewater collection systems.
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Table 5-3. Effect of Corrosion and Odor Control Techniques on VFA Production in Wastewater
Collection Systems
Corrosion/Odor Control Technique
Oxidation of H2S by adding a strong oxidant
(e.g., chlorine, sodium hypochlorite,
hydrogen peroxide)
Application of nitrate salts (common
method)
pH control
Super oxygenation
Vapor-phase odor control
Iron addition (common method)
Effect on VFA Production
• Strong oxidants easily oxidize VFAs.
• Nitrate creates anoxic condition and prevents fermentation, which in
turn prevents VFA formation.
• Nitrate can also act as an electron acceptor (similar to oxygen), so that
oxidizable VFAs are consumed by bacteria.
• pH between 9 and 10 will keep hydrogen sulfide in ionized form and
prevent release to headspace without affecting VFA production.
• pH of 11 will slow or stop the rate of VFA fermentation.
• Excess oxygen added in collection system or in force mains eliminates
production of VFAs.
• Excess oxygen can serve as electron acceptor for heterotrophic bacteria
to uptake VFAs, but without VFA storage.
• Chemical blow down from wet scrubbing has high chlorine
concentration, which can oxidize VFAs.
• Iron addition will precipitate sulfide without impacting VFA formation.
• However, if hydrogen peroxide is added at a downstream location to
reactivate the iron, VFA s will be oxidized.
Source: Derived from Kobylinski et al. (2008).
Industrial discharges to the collection system can significantly impact VFA formation. Discharges
with high temperatures and/or BOD loading can increase fermentation, while discharges containing
nitrate will inhibit VFA production because denitrifying bacteria will use the nitrates to fully metabolize
the organics rather than ferment them. Some industries only discharge seasonally or at certain times
during the week, which can cause great variations in VFAs and rbCOD entering the wastewater plant.
This is more likely to be a problem on weekends when industries shut down. Weekend effects also
occur mostly in suburban bedroom communities.
5.3.2 Anaerobic Fermentation of Primary or Return Activated Sludge
Some treatment configurations, such as the Westbank process, make use of anaerobic
fermentation of the primary sludge to provide VFAs to the BPR process (See Chapter 6 for a detailed
description of this technology). A fermentation process, however, can be added to any configuration to
provide VFAs, especially in areas where little fermentation takes place in the collection system. This
approach has been used very successfully in cold climates.
Fermentation of the primary sludge or the RAS will produce VFA. Primary sludge fermentation is
used more frequently and is generally preferred over fermentation of RAS for reasons provided later in
this section.
In the normal anaerobic fermentation of primary sludge, the larger proteins and carbohydrate
compounds are reduced by acid fermentation to a variety of simple products. Typical fermentation
products are:
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• Acetic acid
• Propionic acid
• Butyric acid
• Hydrogen
Under normal operating conditions at pH near neutral, 60 or more percent of the by-products formed
are acetic acid. A total of 20 to 30 percent could be in the form of propionic and butyric acid with only
small quantities of other fatty acids. VFA production from fermenting primary sludge typically yields
0.066 to 0.15 g VFA/g total solids (both expressed as COD), although values up to 0.3 g VFA/g solids have
been reported (Barnard et al. 2005).
In anaerobic digesters, methane fermentation is established spontaneously following the
biological pathways shown in Figure 5-3. As the redox potential drops, methane organisms will soon
reach an equilibrium wherein the acetates are consumed at the rate at which they are produced. If VFA
production and accumulation is desired the growth of methane organisms should be avoided as much as
possible. This usually can be accomplished by operating the fermenter at SRTs of no more than 4 days.
Acid
Formation
15%
65%
Propionic
Acid
Methane
Formation
Other
Intermediates
Figure 5-3. Biological Pathways of Methane Formation
There are several primary sludge fermenter designs that have been used successfully in full scale
plants. The simplest configuration allows the formation of a thick sludge blanket in the primary clarifier
itself where fermentation takes place. Some of the thickened sludge is returned to the influent of the
primary clarifier to allow for elutriation of the VFA to the primary effluent. This is referred to as an
activated primary sedimentation tank (Barnard 1984). Another variation is to pump sludge to a
complete-mix tank ahead of the primary clarifier to accomplish fermentation. The sludge is then passed
to the primary clarifier for elutriation of the VFA. Both of these processes lead to an increased solids
load on the primary clarifier. Sludge age should be controlled to prevent methanogenic bacteria from
growing and converting the VFA to methane. The SRT of the solids in the fermenter will depend on the
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wastewater temperature which can vary from winter to summer, but usually less than 4 days is
sufficient.
An alternative method accomplishes fermentation in an over-sized gravity sludge thickener by
holding the sludge under anaerobic conditions for typically 4 to 8 days4 (this is referred to as a static
fermenter). The supernatant can then be fed directly to the anaerobic zone, avoiding a high load on the
primary clarifier. Thickening can either be accomplished with a single thickener or in two stages. The
two-stage process can either be a complete mix tank followed by a thickener or two thickeners in series.
It has been shown that adding molasses or other sources of readily biodegradable COD can improve the
performance of fermenters (Bott et al. 2007). Static or two-stage fermenters have the advantage over
fermentation in the primary sedimentation tank because the VFAs in the supernatant can be discharged
directly to the anaerobic zone rather than mixed with the main influent stream. This configuration
allows the by-passing of storm flows without significantly affecting the mass of VFA to the anaerobic
zone. Elutriating water (either from the primary effluent or the final effluent) can be fed to the
fermenter to flush out the VFAs produced and send them to the anaerobic zone.
Some of the RAS can also be fermented in a side stream process at similar SRTs; however, it is
critical that designers carefully consider and account for the potential for phosphorus release. Evidence
from full-scale plants showed that when some RAS or mixed liquor is fermented in an upflow fermenter,
sufficient VFA is produced to overcome the release of phosphorus. RAS or mixed liquor fermentation
could be used in any BPR process, but is most common in processes without primary clarifiers.
Fermenters have several advantages over external carbon sources for BPR, including:
• VFA generation results in a favorable ratio of acetic to propionic acids.
• The VFA produced can enhance denitrification by increasing the growth rate of the
denitrifiers because VFAs are simple, readily biodegradable substrates.
• The size of the anaerobic zone could be reduced to about 5 percent of the total volume.
• The fermenters will supply VFA at a constant rate even through storm flows.
• VFA fermentation for BPR results in smaller increase in sludge production when compared
with precipitating chemicals.
• Fermentation can improve sludge settleability due to selection for PAO's that are good floe
formers.
Disadvantages of fermentation include:
• Increased biological load to the aeration basin.
• Odor issues when compared with chemical removal addition for phosphorus removal (odor
control is typically needed).
4 The actual SRT will depend on the temperature and should be designed to reduce the growth of methanogens.
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• Will reduce digester gas production which can reduce energy output if the plant uses
digester gas to produce electricity.
• Capitol costs of equipment.
See Chapter 10 for recommendations on designing sludge fermentation systems.
5.3.3 Commercial Sources
External sources of VFAs can be added to the WWTP to stimulate BPR. Acetate is typically used
because it is the most efficient VFA. Acetate alone, however, does not match the VFAs produced during
fermentation of municipal sewage. As noted previously in Section 5.2, acetate usually comprises 40 to
60 percent, propionate 20 to 40 percent, and the other VFAs up to 10 percent of total VFAs in municipal
wastewater. Thus, a mixture of acetate and propionate plus small amounts of butyrate and valerate and
their iso-forms more closely represent the products of sewage fermentation. Research has shown that
the mixture significantly affects the competition between PAOs and GAOs. This will be discussed in
greater detail in subsequent paragraphs.
Yuan et al. (2008) examined the effect of the organic carbon source on BPR. The authors
compared the production of PHB (but not the other PHAs) and the uptake and release of phosphorus
achieved using four different sources of organic carbon: acetate, beef extract, glucose, and municipal
wastewater. Experimental results showed that the most effective substrates for phosphorus uptake and
release and PHB production and utilization were small, simple molecules such as acetate and glucose.
Glucose was the substrate that resulted in the most dissolved organic carbon uptake, but that was
partially because it was added in the largest amount. The fastest rate of phosphorus release and highest
uptake occurred when acetate was used as a substrate, as expected. Interestingly, the amount of
phosphorus released per amount of dissolved organic carbon uptake observed was the same for acetate
and municipal wastewater, indicating that the wastewater organics were fermented primarily to
acetate. It should also be mentioned that they acclimated the activated sludge used for the batch tests
using sequencing batch reactors (SBRs) fed by a substrate that was phosphorus limited rather than
organic carbon limited.
Several studies have examined the effects of relative concentrations of acetic and propionic acid
on the growth of PAOs and GAOs. An Australian study shows that while both PAOs and GAOs could use
acetate, PAOs will have a competitive advantage when the VFAs consist of roughly equal parts of acetic
and propionic acid as a growth medium. This is because PAOs that are fed on acetate are able to switch
to propionate much more quickly and effectively than GAOs (Oehmen et al. 2005). This finding led to a
strategy to feed equal amounts of acetic acid and propionic acid as the optimal for stimulating PAO
growth (Oehmen et al. 2006, Bott et al., 2007). Note that studies published by Chen, et al. (2004),
Randall and Chen (2008), and Lopez-Vasquez et al. (2009) have shown that an increase in the propionic
to acetic acid ratio, even if the acetic acid concentration is greater, resulted in higher net phosphorus
removal because the combination favors the PAOs over the GAOs, thereby resulting in higher
concentrations of PAOs in the MLSS.
Another study showed that isovaleric acid drives BPR even better than acetic acid (Bott et al.,
2007). This finding is in contrast to the results obtained by Abu-Ghararah and Randall (1991), who
concluded that isovaleric was the second best VFA for BPR but only 65 percent as efficient as acetic.
Isovaleric acid is much more expensive than acetic acid and is more odorous. It also is not significantly
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generated in the primary sludge fermentation process. Addition of rbCOD such as sugars and alcohols
containing two carbons or more can increase phosphorus uptake by PAOs when added to the anaerobic
zone but may cause sludge bulking if dosed in excess (Jenkins and Harper 2004).
5.4 Environmental Conditions
5.4.1 Dissolved Oxygen and Nitrates in the Anaerobic Zone
When dissolved oxygen is present in the anaerobic zone, heterotrophic aerobic organisms can
use the oxygen as the electron acceptor and will compete with the PAOs for VFAs. This results in less
VFA storage and, subsequently, less BPR. Moreover, if oxygen is available, PAOs can metabolize VFAs
completely and use the energy and carbon obtained for growth instead of storing VFAs using energy
from polyphosphate bonds for subsequent phosphate uptake. If PAOs use oxygen to metabolize VFAs,
they do not release phosphorus in the anaerobic zone during VFA storage, and they do not take up
excess phosphorus in the aerobic zone.
Nitrates can also be used as an electron acceptor by facultative (denitrifying) heterotrophs,
which include some PAOs. The presence of nitrate in the anaerobic zone will deplete the amount of
VFAs available to PAOs, just as the presence of dissolved oxygen does. Nitrates can also inhibit
fermentation of rbCOD and reduce the production of VFAs because most of the fermenting
microorganisms are facultative and can use nitrate as an electron acceptor to fully oxidize the non-VFA
rbCOD instead of producing VFAs as the end product.
Oxygen can be inadvertently added back to the anaerobic zone through recycle flows such as
the RAS and by backmixing from inadequately baffled aerobic zones. Nitrates can be introduced to the
anaerobic zone by the RAS, by mixed liquor recycles from anoxic zones, by backmixing from
inadequately baffled anoxic zones, and by plant recycles such as supernatant from sludge handling
facilities. Significant nitrate interference of BPR is much more common than dissolved oxygen
interference, and efforts to control nitrate introduction to the anaerobic zone should be practiced at all
BPR plants. See section 5.6.2 for additional discussion.
5.4.2 Oxygen in the Aerobic Zone
PAOs need oxygen to digest the storage products and uptake and retain phosphorus in the
aerobic zone. Maintaining a sufficiently high DO transfer in the aerobic zone enhances process stability
and has been found to be a key factor in phosphorus removal (Bott et al. 2007). However, it should be
recognized that the appropriate DO concentration in the mixed liquor for a non-limiting oxygen transfer
rate is a function of the way the biological process is being operated, notably the SRT, the mixed liquor
suspended solids concentration, and the actual HRT in the reactor. For high rate systems, the DO needs
to be 2.0 mg/L or greater in the effluent, i.e., prior to a deoxygenation zone if one is being used. Under
these conditions, phosphorus removal can be improved by increasing the mixed liquor DO
concentration. However, for long residence time systems such as oxidation ditches, excellent
nitrification, denitrification, and excess BPR are possible when DO in the system never exceeds 0.5 mg/L,
and varies from near zero in some sections of the ditch to an effluent concentration of 0.25 mg/L (Sen et
al. 1990). Similar results are possible with other high HRT configurations such as the Schreiber Process.
Re-aeration may be useful to ensure that there is sufficient DO in the MLSS in the final clarifiers to
prevent release of phosphorus during settling.
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5.4.3 pH
Low pH can reduce and even prevent BPR. Below pH 6.9, the process has been shown to decline
in efficiency (WEF and ASCE 2006). Research has shown that it is not possible to establish BPR when the
pH in the anaerobic zone is less than 5.5, even if an abundant amount of VFAs are present in the
anaerobic zone (Tracy and Flammino 1987; Randall and Chapin 1997).
Filipe et al. (2001) found that GAOs do not grow when the pH is greater than 7.25 in the
anaerobic zone. They also showed that low pH values in the aerobic zone will inhibit BPR, but PAOs
have the advantage over the GAOs if the zone pH is 7.0 or greater. Fortunately, robust continuous flow
BPR systems commonly experience a large pH swing between the anaerobic and aerobic zones (e.g.,
from 7.2 ±0.2 to 8.4 or more), and pH control is unlikely to be needed and in fact may be detrimental
when treating municipal wastewaters. Because many wastewater processes such as chemical addition
and nitrification can lower the pH, it should be monitored in the anaerobic zone and chemically adjusted
if necessary.
5.4.4 Temperature
High temperatures (i.e., above 20° C) stimulate GAO acetate uptake rates more than PAO uptake
rates, and temperatures of approximately 30 °C can have an adverse effect on phosphorus removal
(Whang and Park 2006). Bott et al. (2007) have reported that phosphorus removal will generally be
impaired at temperatures greater than 28°C. Modeling studies have shown that GAOs can dominate at
higher temperatures because of their increased ability to uptake acetate at those temperatures
compared to PAOs (Whang et al. 2007).
Panswad et al. (2003) kinetically investigated microbial population dynamics in response to
gradual temperature increases in an enhanced BRP system. As the temperature rose from 20° C to 30° C
to 35.5° C, the predominant microbial group changed from PAOs (47-70 percent of total volatile
suspended solids (VSS)) to GAOs (64-75 percent of total VSS)to the ordinary heterotrophs (90 percent of
total VSS), respectively. Despite the species alteration, the phosphorus contents of the PAOs appeared
to be steady within 0.182 to 0.308 mg/mg VSS (PAO) regardless of the temperature level. The initial
specific phosphorus release rates, which are solely due to the activities of the PAOs, increased with
temperature from between 37.5 and 55.9 to between 51.8 and 61.3, 52.0 and 76.9, 147.2 and 210.3,
and 374.2 and 756.3 mg P/gm VSS (PAO) hr, at 20° C, 25° C, 30° C, 32.5° C, and 35.5° C, respectively.
Although the mean initial specific phosphorus uptake rates of the biomass decreased as the
temperature increased, the data implied that the uptake rate of the PAOs was higher than the other two
microbial groups. These results indicate that the PAOs are lower-range mesophiles or possibly
psychrophiles. As the temperature rises, the portion of energy required for maintenance increases
substantially, which reduces the energy availability for cell reproduction; hence, the PAOs can be
washed out from the system.
Low temperatures can also lower phosphorus uptake, although this has not been an issue in
well operated and properly acclimatized plants (WEF and ASCE 2006). Reduced phosphorus removal
performance in the winter is usually caused by reduced fermentation or increased DO and nitrate inputs
to the anaerobic zone. In general, BPR capacity is greater at temperatures below 15° C than at higher
temperatures if electron acceptor inputs (DO and nitrates) are controlled. In lab experiments, Erdal et
al. (2002) found that PAOs out competed GAOs at 5° C even though the PAO metabolism was slower at
Nutrient Control Design Manual 5-12 August 2010
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5° C than at 20° C. This outcome was because GAOs rely on glycolysis for energy, and glycolysis is very
sensitive to low temperatures, resulting in much slower growth rates for the GAOs. The GAOs virtually
disappeared in the 5° C reactor, resulting in a much greater excess phosphorus removal capacity of the
biomass at 5° C compared to that observed at 20° C because of the much higher concentration of PAOs
in the MLSS.
Although BPR capacity is greater at low temperatures, it has been shown that low temperatures
can cause PAOs to "wash out" of activated sludge before the heterotrophs capable of removing BOD
(McClintock et al. 1991; McClintock et al. 1993; Mamais and Jenkins 1992). Mamais and Jenkins (1992)
showed that the wash out point is determined by a combination of temperature and SRT values. That is,
at each SRT value, there is a temperature that will cause BPR wash out. This has been demonstrated at
both pilot plant and large plant scale.
5.4.5 Cations
The breakdown of polyphosphates increases the phosphate concentration in the cell and the
release of phosphates from the cells. However, because each phosphate molecule (PO4~3) contains
three negative charges, it is unable to pass through the cell membrane on its own. To pass through the
cell membrane, the phosphate molecule must bond with positively charged ions such as magnesium
(Mg+2) and potassium (K+). Once the phosphate molecule bonds with these charged ions, it becomes
neutral and can be transported across the cell membrane. Experiments have shown that magnesium
and potassium are essential cations for BPR rather than just providing charge neutralization, whereas
calcium and other cations that might become involved are not essential (Pattarkine and Randall 1999).
The release of phosphate from PAO cells is a critical step in the anaerobic zone of BPR
wastewater treatment systems. The phosphate will not be released in the anaerobic zone once either
magnesium or potassium is depleted, and BPR will cease. There are no known cases, however, of BPR
limitation because of inadequate magnesium in domestic and municipal wastewaters, but it has been
observed in industrial wastewaters, as well as in the laboratory under experimental conditions.
5.5 Kinetics
The biological processes of WWTPs are designed on the basis of SRT and HRT. The design SRT is
based on the growth rate of the activated sludge microorganisms for the organic loading rate of the
influent wastewater, which determines the oxygen transfer requirements and waste activated sludge
production. The nominal HRT (i.e., the time it takes to fill the reactor(s) if only the wastewater is flowing
into it) is used for reactor design. It determines the mixed liquor suspended solids (MLSS)
concentration for the design SRT value.
5.5.1 Solids Retention Time (SRT)
In general, a minimum system SRT of 3 to 4 days based on total reactor volume is sufficient for
BPR in temperate climate conditions. As long as there are sufficient VFAs available, higher SRTs (as great
as 30 days) will not increase phosphorus uptake because the PAO concentration in the mixed liquor will
increase until either the VFAs or the available phosphates become limiting. If SRT becomes too great,
however, the quality of the effluent will decrease because endogenous respiration will cause release of
phosphorus as biomass degrades. The SRT level at which this secondary release occurs is site specific
Nutrient Control Design Manual 5-13 August 2010
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based on VFA uptake, PHA polymerization and use, and the glycogen breakdown and polymerization
rates in corresponding zones of the system (WEF and ASCE 2006).
5.5.2 Hydraulic Retention Time (HRT)
Both anaerobic and aerobic HRT can impact BPR. In the anaerobic zone, sufficient time is
needed for the fermentation of non-VFA rbCOD to VFAs and for the storage of PHAs. An HRT of 1 to 2
hours is typically required if most of the fermentation needs to take place in the anaerobic zone. VFA
uptake and storage as PHAs, however, is fairly rapid. Thus, the HRT can be much shorter if fermentation
of rbCOD in the anaerobic zone is not a significant source of VFAs (as low as 30 minutes, although one
hour would typically be used). The anaerobic HRT may vary from 5 to 15 percent of the total nominal
HRT of a full biological nutrient removal (BNR) system (removes both nitrogen and phosphorus), with
the actual value depending primarily upon whether or not VFAs are supplied to the anaerobic zone from
a source other than the influent wastewater.
5.6 Important Design and Operational Considerations
This section identifies some key design and operational parameters that can impact BPR. See
Chapter 6 for discussion of specific wastewater plant configurations.
5.6.1 Avoiding Secondary Release of Phosphorus
The release of phosphorus in the anaerobic zone is an essential step of BPR because it indicates
that PAOs are uptaking and storing VFAs as PHBs and other compounds. However, phosphorus release
can occur for other reasons such as low pH, chemical toxicity, and excessive anaerobic respiration
resulting in destruction of PAO cells. Thus, release of phosphorus can also occur in the absence of a
source of VFAs. For example, some of the energy stored as polyphosphate is used for cell maintenance,
and phosphorus is released to the liquid phase. If PAOs release phosphorus without storing polymerized
VFAs, they will not have sufficient energy to remove all of the released phosphorus in the aerobic zone.
This occurrence is known as secondary release.
If PAOs enter the aerobic zone with inadequate PHA, they will not have the energy needed for
complete phosphorus uptake and the efficiency of phosphorus removal will decrease. This may occur in
the following process stages:
• In the anaerobic zone if the HRT is too long and the VFAs are depleted a considerable period
before the end of the retention time.
• In the main anoxic zone when the nitrates are exhausted well before the end of the HRT.
• In the second anoxic zone as shown in Figure 5-4 when there are no nitrates to be removed.
• In the sludge blankets of final clarifiers when the RAS rate is too low and sludge is not removed
fast enough, resulting in the flow of released phosphates over the effluent weir.
Secondary release may also happen in aerobic zones that are too large, resulting in stored
substrate depletion and destruction of PAO cells by endogenous metabolism. Further, excessive
depletion in the aerobic zone of the stored glycogen in the PAOs will reduce VFA storage and
Nutrient Control Design Manual 5-14 August 2010
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phosphorus release in the anaerobic zone, and result in reduced phosphate removal by the BPR system
(Punrattanasin and Randall 2000).
o Anaerobic
O
E
Aeration
I_
o
Q.
O
Q.
o
o
c
_o
1
0)
•*•
0
5 10 15
Hours Retention in Basin
20
Figure 5-4. Example of secondary release in second anoxic zone.
Phosphorus will be released from PAOs in sludge treatment processes that are anaerobic.
Gravity thickening of sludge from BPR processes can lead to phosphorus release if long retention times
are used. Mechanical dewatering instead of gravity dewatering allows less retention time and less
phosphorus release (Bott et al. 2007). Dissolved air flotation (DAF) is usually recommended to thicken
sludge to reduce the amount of phosphorus release. DAF thickening can be quite successful for the
reduction of release, but if the thickened sludge is left on the DAF beach too long before removal, excess
release will occur just as it will when the sludge is left too long in a gravity thickener.
Anaerobic digestion can also lead to extensive phosphorus release. Some of the phosphorus
will, however, be precipitated out as either a metal salt (e.g., calcium phosphate) or as struvite
(magnesium ammonium phosphate, MgNH4PO4). PAOs take up and release magnesium along with
phosphates, and these two ions combine with ammonium (which is always present in abundance in
anaerobic digesters) to form struvite (Mg NH4PO4 • 6H2O). Struvite formation is very fast and will
continue until one of the three ions is reduced to its solubility concentration. Magnesium is usually
present in the lowest concentration, and its depletion typically limits struvite formation within the
anaerobic digester.
Calcium phosphate precipitates also tend to form in anaerobic digesters, but they are non-
stoichiometric and form much more slowly than struvite. If substantial amounts of phosphates are
precipitated by calcium along with the struvite formation, there will be little if any propensity for
struvite to form when the sludge exits the anaerobic digesters. Few struvite problems occur with belt
filter press dewatering of the anaerobically digested sludge, but plants using centrifuges for dewatering
have reported problems with clogging of the ports. Note that if the digested sludge is composted after
dewatering, the resulting Class A sludge will be enriched in magnesium, phosphorus, nitrogen, and, to a
lesser extent, potassium, which also is taken up and released with phosphorus by PAOs. Only thirty
Nutrient Control Design Manual
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percent of the phosphorus entering the anaerobic digesters at the York River plant during four years of
BPR experimentation was recycled back to the headworks from belt filter press dewatering. However, it
was possible to consistently produce an effluent of less than 1 mg/L TP without effluent filtration, in
spite of the phosphorus recycle, following the enactment and implementation of a phosphate detergent
ban in Virginia (Randall et al. 1992).
Alternatives to anaerobic digestion such as composting, drying, or alkaline treatment can be
used to reduce phosphorus release and recycle. There have been several studies that have examined
struvite precipitation as a way to recover phosphorus from digester supernatant. These processes were
tested at full scale facilities in Treviso, Italy and Edmonton, Canada (SCOPE 2004), and currently are still
operating at those two plants and others including plants in Durham, NC, and Olburgen in the
Netherlands.
When anaerobic release of phosphorus occurs, recycling these streams can overload
phosphorus removal processes. The effect can be worsened when the waste handling process is only
operated intermittently. In some instances, there is a high degree of phosphorus precipitation in the
anaerobic digesters, and with sufficient VFAs in the influent, the returned phosphorus may be removed.
However, in many circumstances, some chemicals need to be added to the return streams or to the
anaerobic digester itself so that the metal precipitate will be removed with the dewatered sludge. The
chemical requirement is usually considerably less than when added to the mainstream plant in the
absence of BPR. Removing phosphorus from the recycle stream could also reduce the VFA demand in
the anaerobic zone and in some cases, allow the plant to avoid adding a fermenter or external carbon
source.
5.6.2 Avoiding Backmixing
In configurations where the anaerobic zone is followed immediately by an anoxic or aerobic
zone, backmixing can occur unless the design ensures it is impossible. Backmixing can cause elevated
concentrations of nitrates and/or dissolved oxygen in the anaerobic zone which favors growth of
organisms other than PAOs and thus can inhibit BPR by reducing the VFAs available to the PAOs. In
addition, backmixing can cause thick scum to form on the surface of the anaerobic zone.
Aeration of mixed liquor, either diffused or mechanical, increases the depth of the liquid, which
causes a hydraulic gradient back toward the non-aerated zone. If the baffling is insufficient, flow from
the aerated zone can seep back into the non-aerated (anaerobic or anoxic) zones. Backmixing can be
avoided by increased baffling such as an underflow baffle with openings sized to maintain a forward
velocity of 1 foot per second (0.3 m/sec) or greater at all times or by changing the mixing rates at the
WWTP.
Nutrient Control Design Manual 5-16 August 2010
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5.6.3 Flow and Load Balancing
Flows and loads to WWTPs can vary widely because of regular diurnal flow patterns and because
of larger, more irregular disturbances such as storm events. Peaks in either flow or nutrient load can
stress the BPR system and cause reduced performance as measured by effluent phosphorus
concentrations. Flow peaks can be evened out by using equalization tanks at the head of the plant.
Equalization tanks in combination with nutrient sensors can also be used to balance nutrient loads. In
this case, recycle streams high in nutrient concentrations such as digester supernatant can be stored
during peak nutrient loads and recycled during times when concentrations are low.
5.7 Impacts on Sludge Processing and Handling
Sludges from BPR processes will have higher phosphorus contents, and therefore a higher
settling velocity, but will otherwise be similar to sludge from conventional activated sludge plants. The
stored phosphorus adds dry weight to the sludge and improves settling; however, PAOs produce fewer
VSSthan bacteria that use dissolved oxygen in metabolism because the growth reaction is less efficient.
PAOs can store phosphorus up to 40 percent of their cell mass. Thus, the percent phosphorus in
the wasted activated sludge will be determined by the fraction of the activated sludge that consists of
PAOs. This fraction is determined by the ratio of the VFAs that become available to the PAOs in the
anaerobic zone to the amount of available phosphorus in the anaerobic zone. Either the VFAs or the
phosphorus will become limiting in the anaerobic zone, and the PAO fraction will be determined by
whichever is limited.
BPR for North American sewage usually results in 4 to 8 percent of phosphorus in activated
sludge as a percentage of mixed liquor volatile suspended solids (MLVSS). Phosphorus content as high
as 15 percent of MLVSS, however, was observed for several weeks at the HRSD York River WWTP when
it was operated as an anaerobic/oxic (A/O ) process prior to the phosphate detergent ban in Virginia.
The MLVSS at York River was 10 percent phosphorus for several months during 1986 and 1987. The
effluent soluble phosphorus concentration was 2.2 to 2.5 mg/L during this period of time, illustrating the
inverse relationship between percent P in the MLVSS and effluent TP that is common for EBPR
processes. Thus, for some wastewaters the fraction of phosphorus in the wasted dry solids could be 8
percent or more, but as noted earlier it will commonly be between 4 and 8 percent for North American
sewage. There is a tendency to think that effluent TSS will contain the same concentration of
phosphorus as the MLVSS, but it is much more likely that the effluent TSS from gravity settling will
consist primarily of the low specific gravity floes that contain low concentrations of phosphorus rather
than the high specific gravity PAO floes. Microphotographs consistently show that microbial
distributions in activated sludge are not uniform.
With respect to settling and dewatering, varying results have been found, with some plants
using BPR reporting little or no change in sludge characteristics (Knocke et al. 1992) and others reporting
enhanced settling and dewatering properties (Bott et al. 2007). The sludge produced from BPR will have
higher magnesium and potassium concentrations in addition to phosphorus because the PAOs always
take up these elements with phosphorus during BPR.
As noted in Section 5.6.1, anaerobic digestion of sludge will lead to release of phosphorus into
the liquid recycle stream from dewatering or from supernating of the digester. It is very important that
designers consider this potential when designing for BPR because the phosphorus recycled to the
Nutrient Control Design Manual 5-17 August 2010
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biological process must be removed in addition to the influent phosphorus. However, also as noted in
section 5.6.1, because the resulting digester liquid stream is enriched with ammonium, magnesium and
phosphate, struvite (Mg NH4PO4 • 6H2O) will precipitate in anaerobic digesters until the concentration of
one of the constituents is reduced below its solubility point. If the digester is not supernated (i.e. the
digested sludge is not permitted to settle and form supernatant that is wasted from the system), the
concentrations will remain at high levels and more struvite precipitation will occur than if supernating is
practiced, thereby reducing the amount of phosphorus that will be recycled to the biological process.
With abundant phosphorus and ammonium, magnesium is usually the element in short supply because
the ionic strength of the magnesium taken up and released is only about 50 percent of the ionic strength
of the phosphate uptake and release. Additional precipitation of the released phosphorus with calcium
will usually occur after struvite precipitation is complete. During the HRSD York River study, only about
30% of the phosphate released during anaerobic digestion of the BPR sludge was recycled after
dewatering on a belt filter press (Randall et al. 1992). It was still possible to obtain an effluent TP
concentration of less than 1 mg/L without effluent filtration after the phosphate detergent ban went
into effect. Some designers have proposed chemically precipitating struvite or other phosphate solids to
avoid high phosphorus return in recycle streams (Bott et al. 2007). Depending on where struvite crystals
form, i.e. in the anaerobic digester or after exiting the digester, they can plug centrifuge ports, as well as
pumps and pipes used to convey the sludge, if not controlled. However, because of the higher
concentrations of magnesium and phosphorus in anaerobic digesters when BPR sludge is digested,
most, if not all, of the struvite formation occurs within the sludge and remains with the sludge rather
than forming after the sludge and supernatant exit the digesters. Note, however, the reverse typically
happens when non-BPR sludge is anaerobically digested. The struvite forms after the sludge or
supernatant exit the digester rather than within the digester.
5.8 References
American Public Health Association (APHA), American Water Works Association (AWWA) & Water
Environment Federation (WEF). 2005. Standard Methods for the Examination of Water and
Wastewater, 21st Edition. Eds. A.D. Eaton, L.S. Clesceri, E.W. Rice, A.E. Greenberg, and M.H. Franson.
Abu-Ghararah, Z.H. and C.W. Randall. 1991. The Effect of Organic Compounds on Biological Phosphorus
Removal. Water Science and Technology. 23:585-594.
Barnard, J.L. 2006. Biological Nutrient Removal: Where We Have Been, Where Are We Going? In WEFTEC
2006.
Bott, C.B., S. N. Murthy, T. T. Spano, and C.W. Randall. 2007. WERF Workshop on Nutrient Removal: How
Low Can We Go and What Is Stopping Us from Going Lower? Alexandria, VA: WERF.
Chen, Y., A.A. Randall, and T. McCue. 2004. The Efficiency of Enhanced Biological Phosphorus Removal
from Real Wastewater Affected by Different Ratios of Acetic to Propionic Acid. Water Research. 38(1):
27-36.
Erdal, U.G., Z.K. Erdal, and C.W. Randall. 2002. Effect of Temperature on EBPR System Performance and
Bacterial Community. In Proceedings of WEFTEC 2002.
Nutrient Control Design Manual 5-18 August 2010
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Filipe, C.D.M., G.T. Daigger, and C.P. L Grady Jr. 2001. pH As a Key Factor in the Competition Between
Glycogen Accumulating Organisms and Phosphate Accumulating Organisms. Water Environment
Research. Alexandria, VA: WEF. 73(2): 223-232.
Fuhs, G.W. and M. Chen. 1975. Microbiological Basis of Phosphate Removal in the Activated Sludge
Process for the Treatment of Wastewater. Microbial Ecology. 2(2): 119-38.
Jenkins, D.I. and W.F. Harper. 2004. Use of Enhanced Biological Phosphorus Removal for Treating
Nutrient-Deficient Wastewater. Alexandria, VA: WERF and IWA Publishing.
Knocke, W.R., J.W. Nash, and C.W. Randall. 1992. Conditioning and Dewatering of Anaerobically Digested
BPR Sludge. Journal of Environmental Engineering. 118(5): 642-656.
Kobylinski, E., G.V. Durme, J. Barnard, N. Massart, and S. Koh. 2008. How Biological Phosphorus Removal
is Inhibited by Collection System Corrosion and Odor Control Practices. In WEFTEC 2008.
Lopez-Vazquez, C.M., A. Oehmen, C.M. Hooijmans, D. Brdjanovic, H.J. Gijzen, Z. Yuan, and M.C.M. van
Loosdrecht. 2009. Modeling the PAO-GAO Competition: Effects of Carbon Source, pH and
Temperature. Water Research. 43(2):450-462.
Mamias, D. and D. Jenkins. 1992. The Effects of MCRT and Temperature on Enhanced Biological
Phosphorus Removal. Water Science and Technology. 26(5-6): 955-965.
Mamias, D., D. Jenkins, and P. Pitt. 1993. A Rapid Physical-Chemical Method for the Determination of
Readily Biodegradable Soluble COD in Municipal Wastewater. Water Research. 27(1):195-197.
McClintock, S.A., C.W. Randall, and V.M. Pattarkine. 1991. The Effects of Temperature and Mean Cell
Residence Time on Enhanced Biological Phosphorus Removal. In Proceedings, 1991 Specialty Conference
on Environmental Engineering. American Society of Civil Engineers. New York, N.Y. pp. 319-324.
McClintock, S.A., C.W. Randall, and V.M. Pattarkine. 1993. Effects of Temperature and Mean Cell
Residence Time on Biological Nutrient Removal Processes. Water Environmental Research. 65(2):110-
118.
Melcer, H., P.L Dold, R.M. Jones, C.M. Bye, I. Takacs, H.D. Stensel, A.W. Wilson, P. Sun, and S. Bury. 2003.
Methods for Wastewater Characterization in Activated Sludge Modeling. WERF Final Report. Project
99-WWF-3.
Neethling, J.B., B. Bakke, M. Benisch, A. Gu, H. Stephens, H.D. Stensel, and R. Moore. 2005. Factors
Influencing the Reliability of Enhanced Biological Phosphorus Removal. Alexandria, VA: WERF and IWA
Publishing.
Oehmen, A., A.M. Sanders, M.T. Vives, Z. Yuan, and J. Keller. 2006. Competition between Phosphate and
Glycogen Accumulating Organisms in Enhanced Biological Phosphorus Removal Systems with Acetate
and Propionate Carbon Sources. Journal of Biotechnology. Elsevier Science BV. 123(l):22-32.
Nutrient Control Design Manual 5-19 August 2010
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Oehmen, A., Z. Yuan, L.L Blackall, and J. Keller. 2005. Comparison of Acetate and Propionate Uptake by
Polyphosphate Accumulating Organisms and Glycogen Accumulating Organisms. Biotechnology and
Bioengineering. 91(2). New York, NY: John Wiley & Sons, Inc.
Panswad T., A. Doungchai and J. Anotai. 2003. Temperature Effect on Microbial Community of Enhanced
Biological Phosphorus Removal System. Water Research. 37(2):409-415.
Pattarkine, V.M. and C.W. Randall. 1999. The Requirement of Metal Cations for Enhanced Biological
Phosphorus Removal by Activated Sludge. Water Science and Technology. 40(2): 159-165.
Punrattanasin, W. and C. Randall. 2000. Factors Affecting the Production and Storage of
Polyhydroxyalkanoates in Activated Sludge Biomass. Proceedings, 1st World Congress of the
International Water Association. July 3-7. Paris, France. ISBN: 2-9515416-0-0.
Randall, C. W. and R. W. Chapin. 1997. Acetic Acid Inhibition of Biological Phosphorus Removal. Water
Environment Research. 69(5):955-960.
Randall, A., and Y. Chen. 2008. The Kinetics of Anaerobic and Aerobic Carbon and Phosphorus
Transformation for a Septic Wastewater with Varying Propionate:Acetate Ratios. In WEFTEC 2008.
Randall, C.W., H.D. Stensel, and J.L Barnard. 1992. Design of Activated Sludge Biological Nutrient
Removal Plants. In Design and Retrofit of Wastewater Treatment Plants for Biological Nutrient Removal.
Lancaster, PA: Randall, Ed. Technomic Publishing Co., Inc. pp. 125-126.
SCOPE. 2004. Newsletter No. 57. July. Centre Europeen d'Etudes sur les Polyphosphates. Brussels,
Belgium. Available online: http://www.ceep-
phosphates.org/Files/Newsletter/Scope%20Newsletter%2057%20Struvite%20conference.pdf
Sen, D., C.W. Randall, and T.J. Grizzard. 1990. Biological Nitrogen and Phosphorus Removal in Oxidation
Ditch and High Nitrate Recycle Systems. Report CBP/TRS 47/90. Chesapeake Bay Office, USEPA,
Washington, D.C.
Sperandio, M., V. Urbain, P. Ginestet, M.J. Audic, and E. Paul. Application of COD Fractionation by a New
Combined Technique: Comparison of Various Wastewaters and Sources of Variability. 2001.
Tchobanoglous, G., F. L. Burton, and H.D. Stensel. 2003. Wastewater Engineering: Treatment and Reuse.
New York, NY: McGraw-Hill.
Tracy, K. D. and A. Flammino. 1987. Biochemistry and Energetics of Biological Phosphorus Removal.
Proceeding IAWPRC International Specialized Conference, Biological Phosphorus Removal from
Wastewater. Rome, Italy. September 28-30. In Biological Phosphorus Removal from Wastewater. pp.
15-26. R. Ramadori, Ed. New York, NY: Pergamom Press.
WEFandASCE. 1998. Design of Municipal Wastewater Treatment Plants - MOP 8, 4th Ed. Water
Environment Federation and American Society of Civil Engineers. Alexandria, VA: WEF.
Nutrient Control Design Manual 5-20 August 2010
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WEF and ASCE. 2006. Biological Nutrient Removal (BNR) Operation in Wastewater Treatment Plants -
MOP 29. Water Environment Federation and the American Society of Civil Engineers. Alexandria, VA:
WEFPress.
Whang, LM. and J.K. Park. 2006. Competition Between Polyphosphate-and Glycogen-Accumulating
Organisms in Enhanced-Biological-Phosphorus-Removal Systems: Effect of Temperature and Sludge Age.
Water Environment Research. 78(1): 4-11.
Whang, L.M., C.D.M. Filipe, and J.K. Park. 2007. Model-Based Evaluation of Competition Between
Polyphosphate-and Glycogen-Accumulating Organisms. Water Research. 41(6): 1312-1324.
Xu, S. and S. Hasselblad. 1996. A Simple Biological Method to Estimate the Readily Biodegradable Organic
Matter in Wastewater. Water Research. 30(4):1023.
Yuan, Q., R. Sparling, and J. Oleszkiewicz. 2008. Effect of Different Carbon Sources on Biological
Phosphorus Removal and Polyhydroxyalkanoate Production. In WEFTEC Proceedings 2008.
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6. Overview of Nitrogen and Phosphorus Removal Technologies
Chapter 6 covers:
6.1 Introduction
6.2 Nitrogen Removal Technologies
6.3 Phosphorus Removal Technologies
6.4 Combined Nitrogen and Phosphorus Removal Technologies
6.5 Effluent Filtration
6.6 Sidestream Management
6.7 Technology Performance
6.8 Factors for Simultaneously Achieving Low Nitrogen and Phosphorus
Effluent Concentrations
6.9 References
6.1 Introduction
This chapter describes the many technologies that are available to remove nitrogen and
phosphorus from wastewater. It presents findings from an extensive review of nutrient removal
technologies and techniques currently applied and emerging at municipal wastewater treatment plants
(WWTPs). This chapter also includes a discussion of achievable effluent concentrations based on real-
world data from well-operated plants before ending with a discussion of the key factors in
simultaneously achieving low effluent total nitrogen (TN) and total phosphorus (TP) concentrations.
Refer to Chapters 3, 4, and 5 for detailed discussions on the principles of phosphorus removal by
chemical addition, biological nitrogen removal, and biological phosphorus removal, respectively.
The information presented herein complements findings published in the following EPA reports
on the subject:
• Nutrient Control Design Manual — State of Technology Review Report (Published January, 2009):
Profiles the latest advances in technology to achieve consistently low nutrient levels, including
effluent filtration and advanced clarification techniques, along with up-to-date research on the
removal of emerging microcontaminants. This report was an interim product in the
development of this design manual. Full text available at:
http://www.epa.gov/nrmrl/pubs/600r09012/600r09012.pdf
• Municipal Nutrient Removal Technologies Reference Document (Published September, 2008):
Presents detailed technical and cost information about both biological and physiochemical
treatment technologies for the removal of nitrogen, phosphorus, or a combination of the two.
Includes at least one year's worth of full-scale performance data for 27 wastewater treatment
facilities in the United States and Canada and 9 detailed case studies. Full text available at:
http://www.epa.gov/OWM/mtb/publications.htm
• Emerging Technologies for Wastewater Treatment and In-Plant Wet Weather Management
(Published February, 2008): Describes innovative and emerging technologies for nitrogen and
Nutrient Control Design Manual 6-1 August 2010
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phosphorus removal, assesses their merits and costs, and provides sources for further
technological investigation. Full text available at:
http://www.epa.gov/OWM/mtb/publications.htm
Although this chapter provides some examples of proprietary and emerging technologies, it is
important to recognize that the industry is always changing and that new technologies not identified in
this chapter may emerge in the future. New technologies may be innovative adaptations of existing
technologies or technologies borrowed from another industry, and some could lead to considerable
performance improvements and cost savings. When evaluating new technologies, designers and plant
owners should work closely with their state regulatory agency and use the basic treatment principles in
Chapters 3 through 5 of this design manual to ensure that any new technology follows these basic
principles and can achieve its claimed removal goals.
6.2 Nitrogen Removal Technologies
Nitrification is an aerobic process in which autotrophic bacteria oxidize ammonia or nitrite for
energy production. Ammonia-nitrogen (NH3-N) is first converted to nitrite (NO2~) by ammonia oxidizing
bacteria (AOB). The nitrite produced is then converted to nitrate (NO3~) by nitrite oxidizing bacteria
(NOB). Both reactions usually occur in the same process unit at a WWTP (e.g., activated sludge mixed
liquor or fixed film biofilm). Denitrification is the biological reduction of nitrate or nitrite to nitrogen gas
(N2) in the absence of oxygen (or under "anoxic" conditions). See Chapter 4 for detailed information on
the principles of biological nitrogen removal.
Since nitrogen in treatment plant influents is mostly (70 to 80 percent) ammonia, total nitrogen
removal requires that nitrification occur first followed by denitrification. In the past, some WWTPs were
required only to remove ammonia-nitrogen in wastewater to reduce toxicity to aquatic organisms with
no limits on nitrate or total nitrogen. However, many treatment plants are now required to remove
nitrogen because both ammonia-nitrogen and nitrate-nitrogen can stimulate algae and phytoplankton
growth and lead to eutrophication of U.S. waterways (See Chapter 2 for additional discussion). Many
plants that aren't required to remove nitrogen do so anyway by recycling mixed liquor from the
nitrification aeration tank to an upstream anoxic tank because it produces alkalinity, uses nitrate
produced in aeration zone for BOD removal to thus reduce aeration energy, and improves sludge
settling.
Biological nitrogen removal can be accomplished by a variety of treatment configurations using
suspended growth, attached growth, or combined systems. Nitrification, denitrification and biochemical
oxygen demand (BOD) removal can be accomplished in a single process with bioreactors followed by
secondary clarifiers. Systems can also be designed as separate stage systems with nitrification and BOD
removal occurring in the same bioreactor or in separate bioreactors, and denitrification occurring in a
tertiary process. Membrane bioreactors can be used for solids separation instead of secondary
clarifiers. Physical/chemical methods for nitrogen removal are not commonly used at municipal WWTPs
and are not addressed in this manual.
Sidestream treatment processes can be used to enhance nitrification. Supplemental carbon is
often added for denitrification, and advanced solids separation such as membrane bioreactors (MBR)
and effluent filtration can be used to achieve very low levels.
Nutrient Control Design Manual 6-2 August 2010
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Table 6-1 lists the technologies that are available today (in use and emerging) that can achieve
biological nitrogen removal. Discussion of each technology follows the table. Note that technologies
that achieve both nitrogen and phosphorus removal are discussed in Section 6.4.
Table 6-1. Matrix of Biological Nitrogen Removal Technologies1
Configuration
Single Process Unit
for Nitrification and
Denitrification
Separate Stage —
Nitrification
Separate Stage —
Denitrification
Type
Suspended growth
Attached growth or
hybrid
Suspended growth
Attached growth or
hybrid
Suspended growth
Attached growth
Technology
Modified Ludzack-Ettinger (MLE)
4-Stage Bardenpho
MLE or 4-Stage Bardenpho with Membrane
Bioreactor
Sequencing Batch Reactor (SBR)
Oxidation Ditch with Anoxic Zone
Step Feed Biological Nitrogen Removal
Simultaneous Nitrification Denitrification
(SNdN)
Integrated Fixed Film Activated Sludge
(IFAS)
Moving Bed Biofilm Reactor (MBBR)
Nitrification (See sidestream treatment
processes in Section 6.2.4 for discussion of
nitrification with bioaugmentation)
Biological Aerated Filters (BAF)
Suspended Growth Reactors (not common)
Denitrification Filters
- Downflow
- Upflow Continuous Backwash
Section
Reference
6.2.1.1
6.2.1.2
6.2.1.3
6.2.1.4
6.2.1.5
6.2.1.6
6.2.1.7
6.2.1.8
6.2.1.9
6.2.2.1
6.2.2.2
6.2.3
6.2.3.1
1. These technologies are for nitrogen removal only. Technologies for removal of both nitrogen and phosphorus are presented
in Table 6-3.
6.2.1 Nitrogen Removal in Single Process Unit
This section provides information on technologies that achieve nitrogen removal in a single
sludge process, typically by alternating between anoxic and aerobic conditions. Suspended growth
processes are discussed first followed by fixed film and hybrid systems. Supplemental carbon can be
added to a single unit process to enhance denitrification, although this is more common with separate
stage processes where BOD is depleted prior to the last denitrification step. See Section 6.2.3.3 for a
discussion of the various carbon sources and issues. Sidestream treatment processes to enhance
nitrification are discussed in section 6.6.
6.2.1.1 Modified Ludzack-Ettinger (MLE) Process
The most common nitrogen removal process used at WWTPs, the Modified Ludzack-Ettinger
(MLE) process, is a pre-denitrification, single sludge system. The process includes an initial anoxic zone
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followed by an aerobic zone. In the anoxic zone, nitrate produced in the aerobic zone is reduced to
nitrogen gas using BOD in the anoxic zone influent. Nitrification occurs in the aerobic zone along with
the removal of nearly all remaining soluble BOD. At the end of the aerobic zone, pumps recycle the
nitrate-rich mixed liquor to the anoxic zone for denitrification. See Figure 6-1 for the MLE process
schematic.
Nitrified Recycle
Influent
Anosic
Tint
Aerobic Tank
Secondary
Clarifier
Effluent
WAS
Figure 6-1. Modified Ludzack-Ettinger (MLE) process.
RAS = Return activated sludge; WAS = Waste activated sludge
Source: USEPA 2008b Figure 2-3
The MLE is constructed either by adding an anoxic zone ahead of an aerobic activated sludge
process or by constructing walls in an existing aeration basin to create an anoxic zone at the influent
end, although this will reduce the aeration zone nitrification capacity. A retrofit to an existing activated
sludge plant will require pumping and piping for the internal recycle stream. Nitrogen removal might be
limited by factors such as carbon source availability, process kinetics, and anoxic or aerobic zone sizes.
Oxygen recycled from the aerobic zone can negatively affect the denitrification rate in the anoxic zone.
An additional carbon source may be needed for denitrification.
One challenge of the MLE process and any similar processes that rely heavily on recycling is to
make sure that the system hydraulics are suitable. RAS rates can be quite high and baffles need to be
designed accordingly.
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6.2.1.2 4-Stage Bardenpho
The first two stages of the 4-stage Bardenpho are identical to the MLE system. The third stage is
a secondary anoxic zone to provide denitrification of the portion of the flow that is not recycled to the
primary anoxic zone. The fourth and final zone is a re-aeration zone that serves to strip any nitrogen gas
and increase the dissolved oxygen (DO) concentration before clarification (see Figure 6-2). Methanol or
another carbon source can be added to the third stage to enhance denitrification. Some configurations
have used an oxidation ditch with a low DO zone instead of the first two stages.
The 4-stage Bardenpho process involves several basins and has a larger footprint than the MLE
process. The footprint can be reduced by adding external organic carbon to the second anoxic basin.
Pumping and piping are required for the internal recycle. The process has been combined with MBRs to
achieve low effluent nitrogen.
Nitrified Eec veie
(Optional) Medizno!
Addition
Influent
Tank
Aeration Tank
Aeiobic Tank
.Anoxic
Taut
Secondary
Clarifi-er
Effluent
WAS
Figure 6-2. 4-stage Bardenpho process.
Source: USEPA 2008b Figure 2-5.
6.2.1.3 MLE or 4-Stage Bardenpho with Membrane Bioreactor (MBR)
Membrane bioreactors (MBRs) use membranes placed in the last aerobic zone of either an MLE
or 4-stage Bardenpho treatment system for liquid-solid separation instead of conventional clarification.
Membranes can be submersed in the biological reactor or located in a separate stage or compartment.
See Figure 6-3 for common configuration of a membrane bioreactor in a 4-stage Bardenpho nitrogen
removal system.
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Carbon,
n
TT
^9
1
- ^
? N
i
Waste
•
.1
T3
O
O
CD
CO
^9
0
X
O
c
l^»
1 1 .
r
Sludge
Effluent
Membranes
^j
and Scum
Figure 6-3. Common configuration for a membrane bioreactor 4-stage bardenpho treatment system
Low-pressure membranes (ultrafiltration or microfiltration) are commonly used. Systems can be
pressure or vacuum driven. Despite similar design solids residence times (SRTs) and design principles,
MBR systems operate at a higher mixed liquor suspended solids (MLSS) concentration, which results in
smaller tanks and smaller space requirements than biological nitrogen removal systems with secondary
clarifiers. In addition, membrane separation provides for greatly reduced total suspended solids (TSS) in
the effluent, typically well below 1.0 mg/L, and, hence, slightly greater removal of nitrogen and
phosphorus.
Membrane materials are either organic polymers or inorganic materials such as ceramics. They
are designed in modular units and are typically configured as either hollow fiber bundles or plate
membranes. All membrane systems use an air scour technique to reduce buildup on the membranes
(USEPA 2007b; USEPA 2008a). Membranes require periodic cleaning (typically 2 to 4 times per year)
using citric acid or sodium hypochlorite solution. Operational issues include membrane biofouling,
chemical costs, loss of production for cleaning, and increased pumping requirements with increased
electricity costs (USEPA 2007b; WEF 2005). MBR systems overall produce less waste sludge than
conventional systems because they tend to be operated at higher SRTs. When operated at the same
SRTs there is very little difference, with the MBR system producing slightly more sludge because of
greater effluent TSS capture.
6.2.1.4 Sequencing Batch Reactor (SBR)
Sequencing batch reactors (SBRs) are fill-and-draw batch systems in which all treatment steps
are performed sequentially for a discrete volume of water in a single or set of reactor basins. SBRs use
four basic phases for most systems: fill, react, settle, and decant, followed by an idle period (see Figure
6-4 for a depiction of operating cycles). The SBR control system allows it to mimic the treatment
environments of other suspended growth processes such as the MLE or 4-Stage Bardenpho system by
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use of multiple aeration and anoxic periods. It typically completes 4 to 6 cycles per day per tank when
treating domestic wastewater1.
A common design is two SBRs in parallel to allow treatment of the continuous inflow and for
one to be taken out of service for maintenance, plus one spare basin used for flow storage and
equalization, initially, and becoming the third SBR when expansion is desired. Additional treatment
trains can improve reliability for meeting low effluent goals. The key to the SBR process is the control
system, which consists of a combination of level sensors, timers, and microprocessors, which can be
configured to meet system needs for nitrogen removal (Sen et al. 1990). SBRs are often sold as package
plants and are more commonly used for small community WWTPs. Individual SBR basins are relatively
small because they do not have a separate clarifier, but the overall footprint is typically medium because
designs usually call for multiple SBRs in parallel and the use of an equalization basin. A primary
advantage of SBRs is that settling occurs under quiescent conditions, thus making it more efficient, and
they are easy to automate.
Fill
React
Aeration/mixing
Decant
Figure 6-4. Operating periods of a sequencing batch reactor.
Source: USEPA 2008c.
6.2.1.5 Oxidation Ditch with Anoxic Zone
Oxidation ditches are looped channels that provide continuous circulation of wastewater and
biomass. They typically operate as racetrack configurations around a central barrier, with forward
mixed liquor flows of approximately 1 foot per second or more and long SRTs (e.g., 15 to 30
days),although shorter SRTs are possible. The aerators are typically rotating brushes or turbines that
move the water as well as transfer oxygen. Therefore, no additional pumping or piping is typically
1 SBRs can be operated to achieve biological phosphorus removal (BPR) by using a larger batch reactor with an
anaerobic period; however, chemical addition is more commonly used for treatment to low effluent TP levels
(Young 2008).
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needed compared to a conventional activated sludge system, which reduces energy costs. Oxidation
ditches are widely used in small- to medium-sized communities (WERF 2000a). Oxidation ditches are
commonly designed with a large footprint and thus, are used less for large, urban systems. See Figure 6-
5 for a typical oxidation ditch configuration
Flow
Aerator
WAS
Figure 6-5. Example oxidation ditch configuration.
Source: USEPA 2008b, Figure 2-6
Most oxidation ditches can be operated to remove a substantial amount of TN due to the long
SRTs that ensure complete nitrification, the high MLSS concentrations (often in excess of 3,000 -6,000
mg/L) that encourage simultaneous nitrification-denitrification, and the unique DO profiles along the
flowpath that can be manipulated by controlling oxygen transfer by the rotor mixing systems. Oxidation
ditches that are specifically designed to remove nitrogen typically work by cycling the flow within the
ditch between aerobic and anoxic conditions. TN removal can be enhanced by using automatic controls
(typically DO or Oxidation Reduction Potential) to turn rotors/mixers on and off to maintain desired DO
setpoints along the flow path.
Patented oxidation ditch designs for nitrogen removal include the Carrousel and Phased
Isolation Ditch (PID) (the latter also known as the Biodenitro and Biodenipho) processes. The Carrousel
oxidation ditch is a variation of the traditional design. These ditches are typically deeper than standard
oxidation ditches and use turbine aerators for aeration, mixing, and propulsion instead of rotors. The
PID process uses pairs of ditches operating in alternating anoxic-aerobic or anaerobic-anoxic-aerobic
modes. See section 6.4.1.5 for a discussion of proprietary oxidation ditch designs that can achieve both
nitrogen and phosphorus removal (e.g., the Orbal process)
6.2.1.6 Step Feed Biological Nitrogen Removal
The step feed biological nitrogen removal process splits the influent flow and directs a portion of
it to each of two or more (typically 3 or 4) anoxic-aerobic zone combinations in series with similar
portions of the influent flow going to each zone but a lesser amount to the last anoxic-aerobic
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combination prior to clarification to minimize the discharge of ammonium in the effluent (see Figure 6-
6). The biomass in the later stages is not just treating influent flow but is also reducing nitrate from the
upstream zones.
The step feed system also provides flexibility for systems to handle wet-weather events. The
excessive flow is directed to the last or latter stages and normal flow is maintained in the initial stages.
Step feed systems can be compatible with existing conventional "plug flow" activated sludge processes,
and they do not require the installation of recycle pumps and piping, just step-feed pipes. The footprint
can be large to accommodate both nitrification and denitrification; however, as a retrofit technology, a
step feed system is able to redirect the flows of an existing activated-sludge system without needing to
increase its footprint through adding tank volume. Operational disadvantages include the need to
control the DO concentration of aeration zones preceding the downstream anoxic zones and the need to
control the flow splitting to the step feed points.
Influent
Anoxic
Tank
Aerobic
Tank
Anoxic
Tank
Aerobic
Tank
Anoxic
Tank
Aerobic
Tank
RAS
WAS
Figure 6-6. Step feed biological nitrogen removal.
Source: USEPA 2008b, Figure 2-10
6.2.1.7 Simultaneous Nitrification Denitrification (SNdN)
Simultaneous Nitrification Denitrification (SNdN) is a process whereby DO concentrations are
low enough so that oxygen does not penetrate the entire activated sludge floe. Thus, nitrification is
occurring on the exterior portions of the floe and denitrification is occurring in the anoxic, interior
portions. SNdN commonly occurs in oxidation ditches. SNdN necessitates a larger reactor volume
compared to nitrification only; however, it does not require a separate zone for denitrification and can
result in reduced energy requirements. The need for an additional carbon source for denitrification is
typically reduced or eliminated because the entire process is accomplished in one tank. See Chapter 4,
section 4.14 for additional discussion of SNdN mechanisms.
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One example of a patented technology that uses SNdN is the Schreiber Process. The
wastewater enters a circular basin equipped with a rotating bridge that provides mixing. Aeration is
provided by fine-bubble diffusers attached to the bridge. The aeration can be turned off to sequence
between aerobic and anoxic conditions while the moving bridge continues to keep the tank mixed. The
separation of mixing and aeration makes the system very easy to control for SNdN, as well as for BPR in
the same reactor.
Another technology that takes advantage of SNdN is marketed as the proprietary Symbio®
system by Enviroquip ®, a division of Emico Watertechnologies, Inc (USEPA 2008a). It uses a probe to
measure the level of nicotinamide adenine dinucleotide (NADH) in the biomass and the DO of the
wastewater to predict changes in BOD. Based on instantaneous readings, aeration is adjusted to
maintain optimal conditions (DO < 1 mg/L) for SNdN. Operating data from three municipal WWTPs
show TN removal to 10 mg/L or less (Trivedi and Heinen 2000). This technology is commonly used for
industrial treatment systems.
6.2.1.8 Integrated Fixed Film Activated Sludge (IFAS)
Integrated fixed film activated sludge (IFAS) is a relatively new technology that describes any
suspended growth system that incorporates an attached growth media within the suspended growth
reactor (either aerobic or anoxic zone)2. Many types of fixed and floating media are available, as
summarized in Table 6-2.
Table 6-2. IFAS Media Types, Applications, and Design Considerations
Media Type
Textile or
rope media
Sponge
Plastic
Description
Also called looped-cord or
strand media. Consists of a
polyvinyl chloride-based
material woven into rope or
web with loops along its
length to provide surface area
for the biomass. Arranged
within a rigid frame.
Free floating cuboids with
specific gravity close to water
distributed throughout the
media
Free floating plastic media
Products Names
(Partial List)
AccuWeb
Cleartec
Ringlace
Biomatrix
Linpor
Kaldnes
Hydroxyl iPAC
Entex
Design and Operational Considerations
Rope needs to be located in area with low BOD and
high ammonia concentration for enhanced nitrification.
Middle of aeration zone is usually the optimum
location. Aeration and mixing should provide cross flow
pattern. Worm growth has been reported but can be
controlled by creating anoxic conditions and
chlorinating the RAS. High DO encourages worms, while
low DO increases SNdN and discourages worms.
Requires air knife and airlift pump with an impingement
plate passive squeezing to keep media from sinking.
Requires downstream screen. Replacement of some
media is required during first few years of operation.
Requires downstream screen. Media does not need to
be replaced.
Notes: RAS = return activated sludge; HOPE = high-density polyethylene
Source: USEPA 2008b; WEF and ASCE 2006.
Although most IFAS systems have been used for biological nitrogen removal rather than biological phosphorus
removal, pilot-scale research by Sriwiriyarat and Randall (2005) has demonstrated that BPR can be combined with
IFAS in the same treatment train. Researchers did not use media in the anaerobic zone, and recommend against
this practice. Media was used in the anoxic zone with effectiveness.
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A fine screen is recommended upstream of the secondary process to prevent material such as
hair from interfering with the surface area of the medium (USEPA 2008b). Free floating systems require
installation of effluent screens to keep the media in the reactor (Gellner et al. 2008). The free floating
media also requires more energy input than fixed media because of the need for mixing. Sponges
require recycle as well as mixing.
IFAS has several benefits over suspended growth nitrification systems, including (WEF and ASCE
2006):
• Increased biomass without increased solids loads on the secondary clarifier.
• Can provide more treatment capacity with the same footprint because the selected media is
installed within an existing basin.
• Can improve nitrification under cold temperatures because the attached nitrifiers have a
much longer SRT than the suspended growth ones.
• Sludge has improved settling characteristics.
• Nitrification is more resistance to toxics and washout.
• Low additional operating costs.
Higher residual DO levels are typically needed when non-porous media are used in the aerobic
zone to penetrate the slime layer, which is a feature of this type of media, if the objective is to maximize
nitrification. However, lower DO concentrations can be used to establish simultaneous nitrification-
denitrification. Porous media systems typically operate at residual DO levels of 1.5 to 2.0 mg/L using fine
bubble diffusers. Thomas et al. (2009) found that an IFAS system could be operated with a lower
volume fraction of media if higher DO concentrations were used during low temperature periods.
6.2.1.9 Moving Bed Biofilm Reactor (MBBR)
The moving bed biofilm reactor (MBBR) is similar to the suspended media IFAS systems in that it
uses plastic media with a large surface area to increase biomass within the biological reactor. However,
there is no mixed liquor recycle as there is with IFAS. The MBBR media is submerged in a completely
mixed anoxic or aerobic zone and contained in the zone with screens or sieves. The plastic media are
typically polyethylene with a specific gravity of slightly less than 1.0. The carrier elements from most
manufacturers are shaped like cylinders or wheels with internal and external fins. These shapes provide
a high surface area per unit volume that is protected from shear forces, allowing better biofilm growth
(USEPA 2008b). Slow speed submersible mixers are typically used in anoxic zones, while aeration is
typically supplied by coarse bubble diffusers in aerated zones (WEF and ASCE 2006) to obtain good
mixing.
Like IFAS, MBBR can be used in separate aerobic and anoxic zones. They can reduce solids
loading, generate sludge with better settling characteristics, and prevent inhibition and washout of
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nitrifiers in aerobic zones. Unlike IFAS, MBBR systems do not involve any return flow and do not rely on
suspended growth to provide treatment.
6.2.2 Separate Stage Processes—Nitrification
6.2.2.1 Suspended Growth Nitrification
Suspended growth nitrification and BOD removal are commonly accomplished in the same
aerated basin as long as two conditions exist: (1) the biomass inventory is retained long enough to
establish a stable population of nitrifiers and (2) the hydraulic retention time (HRT) is such that the
biomass can react with the ammonia-nitrogen entering the system. Because nitrifying bacteria have
slower growth rates than BOD-consuming bacteria, nitrification kinetics govern the process design, with
the critical design parameter being design SRT.
With sufficient DO (generally 2 mg/L or greater) and adequate pH (generally at least 6.8), typical
design SRTs required for nitrification range from 10 to 20 days at 10°C and 4 to 7 days at 20°C (Randall et
al., 1992). The nitrification kinetics and thus, design SRT, vary considerably from system to system
depending on site-specific factors such as operating DO, mixing, floe size, temperature, alkalinity, and
presence of inhibitors in the influent. Because of this variability, designers are strongly encouraged to
determine site-specific nitrifier growth rates rather than using values from industry literature. See
Chapter 10 for guidelines on determining site-specific nitrifier growth rates. See Chapter 4 for an in-
depth discussion on the microbiology of nitrification, reactions and stoichiometry, nitrification kinetics,
and factors affecting nitrifier performance at WWTPs.
The SRT needed for near complete nitrification of influent ammonia can be reduced significantly
by using a series of aeration tanks (staged systems) versus a single aeration tank. The effect is less
pronounced for systems with high return sludge recycle rates, such as MBRs, in which the influent
nitrogen is substantially diluted. For plants having difficulty in nitrifying due to insufficient SRT, there
are some emerging sidestream processes that can increase the nitrification rate. One of these is
bioaugmentation. Bioaugmentation is accomplished by seeding the activated sludge process with an
external source of nitrifying bacteria (also known as external bioaugmentation) or making process
improvements to increase the activity of or enrich the nitrifier population (also known as in situ
bioaugmentation). See Section 6.6 for a discussion of several bioaugmentation technologies.
6.2.2.2 Attached Growth Nitrification
Trickling filters and rotating biological contactors (RBCs) have historically been used for
biological treatment of wastewater and can achieve nitrification with a low organic loading and a
relatively high media volume. Typically, nitrification is achieved on the media after most of the BOD is
removed because the heterotrophic population competes with the nitrifying organisms for oxygen and
space on the media. Major disadvantages of these technologies compared to suspended growth
systems is that the SRTs cannot be controlled, and denitrification is either fully dependent on the
addition of a supplemental carbon source or a two stage system must be constructed, e.g. anoxic filter
followed by an aerobic filter with recycle of nitrates back to the anoxic filter. Suspended growth
processes, on the other hand, are more flexible because the SRT can be precisely controlled, and they
can be designed to denitrify 80 percent or more of the nitrate resulting from near complete nitrification
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using the incoming BOD as the carbon source, which is a lower cost solution. Consequently, trickling
filters and RBCs have fallen out of favor for nutrient removal applications.
Existing trickling filters, however, can be integrated into a BNR system (Hu 2001). One approach
it to first treat wastewater in an anaerobic activated sludge basin, followed by internal settling of the
mixed liquor. The clarified liquid is sent to the trickling filter for nitrification, whereas the settled flow is
routed to the main anoxic tank, which also receives the nitrified effluent from the trickling filter, where
denitrification is accomplished. The denitrified flow from the anoxic zone goes to a small aerobic basin
to strip the nitrogen gas out of the flow prior to settling the mixed liquor in an anoxic basin. From the
anoxic basin, the flow is briefly aerated to strip the nitrogen gas and the flow is settled with the RAS
returned to the anaerobic zone. This can be accomplished with an activated sludge SRT of 8 to 10 days
in temperate climates and also is useful for controlling filamentous microorganisms.
Newer attached growth systems called biological aerated filters (BAF) have taken the place of
trickling filters and RBCs. They differ from trickling filters in that the media is smaller, i.e. has a higher
specific surface area, and a built-in aeration system. BAFs originated in Europe and now are installed in
several locations in North America (Sen et al. 2008). BAFs can be configured in series to remove BOD in
one unit and oxidize ammonia-nitrogen in the next, or they can be designed for BOD removal and
nitrification in a single unit, depending on process goals. Additionally, they are used for denitrification.
BAFs require higher DO concentrations in the bulk water compared to suspended growth nitrification to
promote DO diffusion into the biofilm. Vidal et al. (1997) recommend a minimum DO concentration of
at least 6 mg/L; recent pilot- and full-scale studies in California found good nitrification at a consistent
minimum DO concentration of 7 mg/L (Holloway et al. 2008). Advantages of BAFs include their smaller
footprint, higher hydraulic loading rate, and less susceptibility to washout than suspended sludge
systems (Verma et al. 2006). BAF performance is also relatively insensitive to temperature (Zhu and
Chen 2002; Holloway et al. 2008), which can be a significant advantage in cold weather regions. BAF
technology can also be configured for pre- or post-denitrification, with post-denitrification requiring the
addition of a supplemental carbon source.
The EPA emerging technologies report (USEPA 2008a) identified two proprietary BAF designs as
established technologies: the Biofor® system and the Biostyr® system. The Biofor® filtration system is a
fixed bed, upflow, expanded bed system with dense granular clay media. Air is sparged into the filter to
maintain an aerobic environment. The Biostyr® system is similar but uses media that are less dense
than water, and the system operates as an upflow packed bed held in place during operation by a screen
at the top of the cell.
6.2.3 Separate Stage Processes—Denitrification
A separate-stage denitrification system may be appropriate for plants that are regularly
achieving nitrification and need to add denitrification capabilities. Suspended growth systems are not
common, although they have been used for some treatment plants. Suspended growth reactors
typically have short SRTs and a small aerated zone following the denitrification zone to oxidize excess
methanol and release nitrogen gas bubbles contained in the denitrified mixed liquor (WEF and ASCE
2006).
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6.2.3.1 Denitrification Filters
Denitrification filters were first installed in the 1970s and are a popular add-on technology for
wastewater utilities trying to consistently achieve low effluent TN limits. They have the advantage of
providing both denitrification and effluent filtration. Because the BOD has been removed from the
wastewater during secondary treatment, a supplemental carbon source must be added (see Chapter 4
for an extensive discussion on consumption rates and kinetics of exogenous carbon sources).
Denitrification filters typically have a small footprint compare to attached growth systems, but require
additional pumping and energy costs.
Denitrification filters have evolved into two main process configurations: downflow filters and
upflow continuous backwash filters. Downflow denitrification filters are deep bed filters consisting of
media, support gravel, and a block underdrain system. Media depths are typically 8 to 12 feet, and
loading rates are between 2 and 3 gallons per minute per square foot (gpm/ft2). Wastewater flow is
directed over weirs onto the top of the filter where a supplemental carbon source, typically methanol, is
added. Backwashing (typically air scouring and backwashing with air and water) is conducted at regular
intervals to remove entrapped solids from the filter. During operation, nitrate is converted to nitrogen
gas and becomes entrained in the filter media, increasing head loss through the filter. To release
entrained nitrogen, most downflow denitrification systems have a nitrogen-release cycle operation that
essentially "bumps" the filter by turning on the backwash pump(s) for a short period of time. See Figure
6-7 for a schematic of a typical downflow denitrification filter.
Filter Backwash
Influent
1 *•
Secondary Treatment
Process
Secondary
Clarifier
(if used)
RAS(ifused)
Methaiiol or other
carbon source
Denitrifying
Filter
Effluent
WAS
Figure 6-7. Downflow denitrification filter.
Source: USEPA (2008b), Figure 2-1
Upflow continuous backwash filters have the advantage of remaining in service during
backwashing, as they are an integral part of the filtering process. Wastewater enters the bottom of the
filter where a carbon source, typically methanol, is added. Water flows up through an influent pipe and
is dispersed into the filter media through distributors. Filtered water discharges at the top of the filter.
Filter media continuously travel downward, are drawn into an airlift pipe at the center of the filter, and
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are scoured before being returned to the filter bed. See Figures 6-8 and 6-9 for process schematics of
upflow continuous backwash filters.
Inlet Water
Reject/
washwater
'V — Filtrate Outlet
Water
Sandwasher
Air-lift pump
Water
Distributer
Drain
Figure 6-8. Continuous backwash upflow sand (CBUS) filters.
Source: Feldthusen 2004. © Nordic Water Products AB. Used with permission
Figure 6-9. Close-up of continuous backwash upflow sand (CBUS) filter.
Source: Felduthsen 2004. © Nordic Water Products AB. Used with permission
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Performance of denitrifying filters depends on many factors including:
• Influent weir configuration (needs to reduce DO, which can inhibit denitrification at
concentrations as low as 0.2 mg/L)
• Filter media
• Underdrain system
• Backwash system
• Flow and methanol feed control to avoid increased BOD in the plant effluent
Operators at a wastewater system in Connecticut reported that key issues for them were influent piping
design to minimize aeration, maintaining a consistent flow to the filters, and control of methanol feed
based on influent chemical oxygen demand (COD) (Pearson et al. 2008).
6.3 Phosphorus Removal Technologies
WWTPs remove phosphorus by chemical or biological methods. A combination of these
methods (e.g., biological phosphorus removal and chemical polishing) can be used to achieve very low
effluent TP levels at minimal costs. Table 6-3 lists the chemical and biological methods for phosphorus
removal. A discussion of each technology follows the table.
Table 6-3. Matrix of Phosphorus Removal Technologies1
Configuration
Chemical Precipitation
Followed by Solids
Separation
Biological
Type
Pre-, co-, or post-
precipitation
Suspended growth
Technology
Addition of lime or metal salt at one or multiple
locations within the plant to react with phosphorus
to form precipitate. Solids can be removed through
conventional clarification or advanced solids
separation processes. Tertiary filtration can remove
additional solids and produce effluent with very low
TP concentrations. (See Section 6.5 for discussion of
tertiary filtration.)
Anaerobic/Oxic (A/0), i.e. Phoredox
Oxidation Ditch with Anaerobic Zone
Section
Reference
6.3.1
6.3.2.1
6.3.2.2
1. These technologies are for phosphorus removal only. Technologies for removal of both nitrogen and phosphorus are
presented in Table 6-4 and discussed in Section 6.4.
6.3.1 Phosphorus Removal by Chemical Addition
Chemical precipitation for phosphorus removal is a reliable, time-tested wastewater treatment
method that has not drastically changed over the years. To achieve removal, various chemicals are
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added to wastewater where they react with soluble phosphates to form precipitates. The precipitates
are removed using a solids separation process, most commonly settling (clarification). Chemical
precipitation is typically accomplished using either lime or a metal salt such as aluminum sulfate (alum)
or ferric chloride. The addition of polymers and other substances can further enhance floe formation
and solids settling. Operators can use existing secondary clarifiers or retrofit primary clarifiers for their
specific purposes. See Chapter 3 for a detailed discussion on the principles of phosphorus removal by
chemical addition.
The effectiveness of phosphorus removal by chemical addition is highly dependent on the solids
separation process following chemical precipitation. The direct addition of metal salts to activated
sludge processes followed by conventional clarification can typically remove TP to effluent levels
between 0.5 and 1.0 mg/L (Bott et al. 2007). Tertiary processes (post-secondary treatment) can be used
to remove phosphorus to very low concentrations. For example, Reardon (2005) reports that four
WWTPs with tertiary clarifiers achieved TP levels of between 0.032 and 0.62 mg/L.
Two common tertiary processes are clarification and effluent filtration. These approaches can
be used separately or in combination. Section 6.5 describes effluent filtration technologies.
Conventional clarification and recent advances in tertiary clarification processes are discussed below.
Conventional Clarification
Conventional gravity clarifiers can be used very effectively for the tertiary removal of
phosphorus precipitates. Their performance can be enhanced by the addition of synthetic polymers
following precipitation to aid flocculation and sedimentation. The precipitating chemical typically would
be added to the flow upstream of the clarifier, and polymer commonly would be added at the center
well of the clarifier. This technology following simultaneous precipitation and clarification in the
activated sludge process was used to obtain effluent TP concentrations of less than 0.18 mg/L (monthly
average) at the Fairfax County, VA, Lower Potomac WWTP and the Arlington County, VA, WWTP for
more than two decades. The tertiary clarifiers were followed by sand filtration.
Lime clarification, although not commonly used today, is a viable treatment option for tertiary
removal of phosphorus (WEF 2005). Most existing plants use a two-step process whereby excess lime is
added to the first stage to raise the pH to between 11 and 12. The effluent is passed to a second stage
where carbon dioxide and possibly soda ash is added to stabilize the pH to between 9.5 and 10.5 and to
precipitate additional calcium (re-carbonation). Sand filtration is used after re-carbonation. This
technology has been used at the Upper Occoquan Sewage Authority Advanced Wastewater Treatment
Plant, Fairfax County, VA, since 1978 to achieve effluent TP concentrations of less than 0.1 mg/L (weekly
average).
Conventional clarification has been used for many years and design and operation is well
covered in other manuals such as the WEF manual of practice for design of municipals WWTPs (WEF and
ASCE 2010) and the WEF Clarifier Design Manual (WEF 2005).
High-Rate Tertiary Clarification Processes
Two types of high-rate clarification processes are common: dense sludge and ballasted high-
rate clarification (BHRC). The latter has been used successfully to achieve low effluent phosphorus
concentrations (WEF 2005). Several patented BHRC technologies using different types of ballast, such as
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recycled sludge, microsand, and magnetic ballast (USEPA 2008a), have been developed in recent years.
The advantages of high rate clarification are that the clarifiers have a smaller footprint and are able to
treat larger quantities of wastewater in a shorter period of time. The following patented processes are
examples of high rate clarification (including performance estimates):
• DensaDeg® uses a coagulant in a rapid mix basin to destabilize suspended solids. The water
flows into a second tank where polymer (for aiding flocculation) and sludge are added. The
sludge acts as the "seed" for formation of high density floe. This floe is removed in settling
tubes (USEPA, 2008). The main advantages of this process are a smaller footprint and
denser sludge, which is easier to dewater. Pilot testing for the City of Fort Worth, TX found a
phosphorus removal rate of 88-95 percent for DensaDeg® (USEPA 2003). See Figure 6-10 for
the process configuration for this technology.
Coagulation
Flocculation
Clarification
Sludge Recirculation
Figure 6-10. DensaDeg® high rate clarification process flow diagram.
Source: USEPA 2008a
Thickened Sludge
Extraction
• Actiflo® uses a coagulant in a rapid mix basin to destabilize suspended solids. The water
flows to a second tank where polymer (for aiding flocculation) and microsand are added.
Microsand provides a large surface onto which suspended solids attach, creating a dense
floe that settles out quickly. Clarification is assisted by lamella plate settling. Product pilot
testing in Fort Worth, TX showed a phosphorus removal efficiency of 92-96 percent for
Actiflo® (USEPA 2003).
• The CoMag™ process uses the addition of high density magnetite ballast with metal salts
and polymer to promote floe formation and high rate settling. Settling is followed by
magnetic separation for final effluent polishing and recovery of the magnetite ballast from
the tertiary sludge, utilizing its strong natural attraction to magnetic fields (USEPA 2008a).
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Since November 2007, CoMag™ has been in operation at a 1.2 million gallon per day (MGD)
WWTP in Concord, MA (EPA Region 10 2007; Proakis Ellis and Cathcart 2008). Shortly after
startup, the system reached 0.05 mg/L TP in the plant effluent on an average daily basis and
has performed to this level consistently under a wide range of influent loading conditions.
Proakis Ellis and Cathcart (2008) report many operational advantages of CoMag™ including
reduced chemical usage and less sludge production compared to using alum alone for
secondary treatment. The final polishing electromagnet in Concord requires relatively high
power usage; however, newer installations of the CoMag™ system are being designed with
final polishing magnets that do not require any power usage. See Figure 6-11 for the
process configuration for this technology.
Polymer J^>—
> pH Contro[
Metal Salt
\ Secondary's.
/ Effluent /
Magnetic
Ballast
Figure 6-11. CoMag process flow diagram.
Source: USEPA 2008a
6.3.2 Biological Phosphorus Removal
Biological phosphorus removal (BPR) is accomplished by encouraging the growth of phosphate
accumulating organisms (PAOs). Under anaerobic conditions, PAOs uptake and store simple carbon
food sources such as volatile fatty acids (VFAs) using the energy in phosphate bonds, and release
phosphorus to solution. When the PAOs are subject to aerobic conditions, they metabolize the stored
carbon to generate energy for cell growth and maintenance and store excess energy by taking up
phosphate ions and creating polyphosphates. The phosphorus uptake by PAOs in the aerobic zone
results in a net reduction in phosphorus in the wastewater when sludge is wasted.
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The success of biological phosphorus removal is highly dependent on the availability of VFAs in
the anaerobic zone. VFAs are produced by fermentation of municipal wastewater, which can occur in
the collection system or in the anaerobic zone itself. If sufficient VFAs are not present, plant operators
can consider adding VFAs by using onsite sludge fermentation or by adding a commercial source. See
Chapter 5 for additional information on the principles of BPR.
All biological phosphorus removal technologies are designed to cycle treatment from anaerobic
to aerobic conditions to encourage PAOs to grow and uptake phosphorus. This section describes two
technologies that have been designed to biologically remove phosphorus only (not nitrogen) from
wastewater. Most technologies have been designed to remove both nitrogen and phosphorus and are
presented in Section 6.4.
As noted in Chapter 5, nitrate in the recycle stream can inhibit BPR. This can be a problem for
plants that nitrify but do not denitrify. In these cases, nitrification needs to be eliminated or greatly
reduced to biologically remove phosphorus unless the readily biodegradable COD (rbCOD) in the
influent is very high, which it typically will not be for municipal and domestic wastewaters. Operating at
an SRT less than 4 days will usually washout nitrifiers at temperatures less than 25 °C. Note that most
WWTPs in the Eastern USA have effluent ammonia limits, which means that operating without
nitrification is frequently not possible. Once a plant is practicing nitrification, there are additional
advantages to also practicing denitrification for TN removal (see Section 6.8).
6.3.2.1 Pho-redox (A/O)
The Pho-redox (A/O) process is a conventional activated sludge system with an anaerobic zone
at the head of the aeration basin. The RAS is pumped from the clarifier to the anaerobic zone. It is a low
SRT process that is operated to avoid nitrification. With no nitrates in the RAS, the process is reliable
and easy to operate except at mixed liquor temperatures in excess of 25°C when nitrification is difficult
to avoid. If nitrates are present in the recycle stream, the anaerobic zone can be split into an anoxic
chamber for nitrate denitrification and one or more anaerobic zones for biological phosphorus removal.
Figure 6-12 provides a schematic for this system.
The technology is relatively easy to retrofit into an existing basin by installing a baffle wall and
mixers to produce an anaerobic zone. The aerators will need to be redistributed in the aerobic zone,
with closer spacing in the influent end of the zone.
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Influent
Aoaarobk
Aerobic Tsui
Seeoodaiy
Cfajifier
Effluent
RAS
WAS
Figure 6-12. Pho-redox process (A/0)
Source: USEPA 2008b, Figure 2-15
6.3.2.2 Oxidation Ditch with Anaerobic Zone
Similar in design to the oxidation ditches for nitrogen removal discussed previously, oxidation
ditches for phosphorus removal are a modification of the standard design with the requirement that an
anaerobic zone be established. The anaerobic zone is most often ahead of the ditch (sometimes called a
"selector") or it can be within the ditch if the DO balance is carefully managed, but filamentous
organisms will tend to grow. See Figure 6-13 for an example of an oxidation ditch with an anaerobic
basin (also called a "selector").
Oxidation ditches with phosphorus removal but not nitrogen removal may encounter problems
with low alkalinity because nitrification will deplete alkalinity but denitrification in an anoxic zone is not
present to partially replenish it. Lower TP effluent levels can be achieved by close monitoring and
control of DO and flow in the anaerobic zone. The footprint size for this process is potentially large
compared to other technologies. It does not require any additional recycle pipes or pumping and needs
minimal energy to operate.
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Anaerobic
Zone
(Selector)
Flow-
Aerators
Influent
II
I
II
Secondary
Claiifier
RAS
WAS
Effluent
Figure 6-13. Oxidation ditch with anaerobic zone.
Source: USEPA (2008b), Figure 2-16
6.4 Combined Nitrogen and Phosphorus Removal Technologies
Table 6-4 presents the technologies capable of simultaneously achieving both nitrogen and
phosphorus removal. A discussion of each technology follows the table.
Table 6-4. Matrix of Combined Biological Phosphorus and Nitrogen Removal Technologies
Configuration
Biological
Hybrid Chemical /
Biological
Type
Suspended growth
Technology
3 Stage Pho-redox (A2/0)
5-Stage Bardenpho
University of Capetown (UCT),
Modified UCT, and Virginia Initiative
Project (VIP)
Westbank Configuration
Oxidation Ditch with Anoxic and
Anaerobic Zone
Sequencing Batch Reactor (SBR)
Blue Plains Process
Biological Chemical Phosphorus and
Nitrogen Removal (BCFS)
Section Reference
6.4.1.1
6.4.1.2
6.4.1.3
6.4.1.4
6.4.1.5
6.4.1.6
6.4.2.1
6.4.2.2
6.4.1 Biological
6.4.1.1 3 Stage Pho-redox (A2/0)
The 3 Stage Pho-redox (A2/O) process adds an anoxic zone to the A/O process configuration
after the anaerobic zone to achieve denitrification. In addition, nitrate-rich liquor is recycled from the
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end of the aerobic zone to the head of the anoxic zone to enhance denitrification, as shown in Figure 6-
14. A shortcoming of the 3 Stage Pho-redox process is that there will be nitrates present in the RAS,
potentially making the phosphorus removal process unreliable. With A2/O, as with all combined
nitrogen-phosphorus biological systems, some phosphorus is taken up in the anoxic zone by the PAOs,
and the sludge residence time in each zone must be sufficient to allow complete phosphate release or
uptake. The clarifier also must be operated to regularly waste solids to avoid release of phosphate by
the endogenous respiration of PAOs.
Nitrified Recvcle
Anaerobic
Tank
Anoxjc
Tank
Aerobic Tank
Secondary
Clmfier
Effluent
RAS
WAS
Figure 6-14. 3 Stage Pho-redox process (A2/0).
Source: USEPA 2008b, Figure 2-22
6.4.1.2 5-Stage Bardenpho
The 5-stage Bardenpho process consists of the 4-stage process (see 6.2.1.2) with an anaerobic
zone added to the front of the system (see Figure 6-15). A nitrate-rich liquor is recycled from the first
aerobic stage to the first anoxic zone. The RAS is recycled from the clarifier to the beginning of the
anaerobic zone. Since the nitrates in the RAS are typically low (from 1 to 3 mg/L), they do not have the
potential to significantly interfere with the phosphorus removal process as with the A2/O configuration
(see 6.4.1.1). Methanol might need to be added to the second anoxic zone for complete denitrification
or to minimize the volume of the second anoxic zone.
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Nitrified Recycle
Anaerobic Tank
Influent
(Optional) Methanol
Addition
Aerobic Tank
Aiioxic
Tank
Aerobic Tank
Aiioxic
Tank
Secondary
Clarifier
Effluent
RAS
WAS
Figure 6-15. 5-stage Bardenpho process. Note that alternative external carbon sources can be used instead of methanol
Source: USEPA (2008b), Figure 2-23
6.4.1.3 University of Capetown (UCT), Modified UCT, and Virginia Initiative Plant (VIP)
The UCT process was designed to reduce nitrates to the anaerobic zone when high removal of
nitrates in the effluent is not required. Similar to other nitrogen and phosphorus removal technologies,
it consists of three stages: an anaerobic stage, an anoxic stage, and an aerobic stage. The RAS is
returned from the clarifierto the anoxic zone instead of the anaerobic zone to allow for denitrification
and to avoid interference from nitrate with the activation of the PAOs in the anaerobic stage. A nitrate
rich stream is recycled from the aerobic zone to the anoxic zone influent. Denitrified mixed liquor is
recycled from the anoxic zone effluent to the anaerobic zone.
The modified UCT process splits the anoxic zone into two stages. The nitrate rich recycle from
the aerobic zone is recycled to the head of the second anoxic stage. The distinction between the
modified and the UCT processes is seen in Figure 6-16. The nitrate containing RAS is recycled to the first
anoxic stage where it is denitrified. Next, the denitrified RAS is recycled from the end of the first anoxic
stage back to the head of the anaerobic stage and mixed with the incoming wastewater.
The Virginia Initiative Plant (VIP) is similar to the UCT process, but the anaerobic and anoxic
zones are baffled into two or more sections each to increase rates of reaction in the first section of each
zone, thereby firmly establishing the desired anaerobic and anoxic conditions in the second section.
These configurations have a medium-sized footprint and, in some cases, can be installed in
existing basins. They require extensive pumping, piping, and control for the multiple recycle streams.
However, internal recycles can be inexpensively installed when retrofitting most activated sludge basins
by using low-head pumps, such as propeller pumps, and piping along one internal wall of the basin.
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Minified Recvcie
Mixed Liquor
Recycle
WAS
UCT Process
Nitrified Recycle
Influent
Anaerobic
Tank
Anoxic
Tari
^J
Amoxic
Tank
Aerobic Tank
Secondary
Ckiifier
n
HAS
WAS
Modified UCT Process
Figure 6-16. UCT and Modified UCT process.
Source: USEPA 2008b, Figures 2-24 and 2-25
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6.4.1.4 Westbank
The Westbank process, shown in Figure 6-17, has a small, pre-anoxic zone followed by an
anaerobic zone, a second anoxic zone, and lastly, a large aerobic zone. The pre-anoxic zone minimizes
DO and nitrates entering the anaerobic portion. Primary effluent is divided among the pre-anoxic zone
(to denitrify the RAS), the anaerobic zone (to provide some VFAs for phosphorus removal), and the
second anoxic zone (to stimulate denitrification). Direct feed of primary effluent to the second anoxic
zone increases the denitrification rate and therefore decreases the size of the anoxic zone compared to
the 5-stage Bardenpho system. During storm flows, excess flow is passed directly to the main anoxic
zone. VFAs obtained from fermentation of the primary sludge are passed to the anaerobic zone.
Additional piping is required to discharge the influent to multiple zones; however, recycle
streams are reduced compared to similar layouts. The fermenter requires additional pumping, piping,
and controls for optimum fermentation and feeding of the VFAs to the anaerobic zone. The design has a
medium-sized footprint.
Effluent
KJ
Anoxic J Anaerobic Aerobic
Figure 6-17. Westbank process.
6.4.1.5 Oxidation Ditch with Anoxic and Anaerobic Zones
Section 6.2.1.5 described oxidation ditches that are biologically configured to remove nitrogen.
Any of these configurations can be modified with an anaerobic basin (also called a "selector") before the
ditch to accomplish BPR in a combined system. Because of the very high internal recycle within the
ditch, very low nitrate concentrations can be achieved in the mixed liquor before settling, which
minimizes nitrate in the RAS stream to the anaerobic selector zone, improving BPR process efficiency.
Many oxidation ditch configurations have been developed to simultaneously achieve biological
nitrogen and phosphorus removal to low levels. An example is the VT2 process developed at Bowie,
MD, which operates two Pasveer ditches in series with dedicated anoxic, near anaerobic, and aerobic
zones (See Figure 6-18). It also has a sidestream anaerobic zone that receives approximately 30 percent
of the influent to enhance BPR. Denitrified MLSS for the anaerobic zone are obtained from the end of
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the near anaerobic zone of the adjacent ditch. Operated without primary sedimentation, the system
consistently achieved very low annual average effluent TP concentrations (0.16 - 0.24 mg/L) without
chemical addition or effluent filtration for more than 20 years. Originally designed for only one ditch,
both ditches had to be operated in series because the plant has limited clarification capacity. Series
operation of both ditches resulted in lower MLSS concentrations to the clarifiers. The Biodenipho
process uses pairs of ditches that operate in alternating anoxic-aerobic modes. An anaerobic tank with
RAS returned to it is placed before the ditches to stimulate BPR, and the ditches are alternated between
aerobic and anoxic conditions for nitrification and denitrification. The Orbal process uses anaerobic and
anoxic zones in the outer of three concentric oval-shaped ditches with RAS recycled from the clarifier to
the anoxic zone to remove both nitrogen and phosphorus.
TP< 0.25 mg/L
TN < 4.0 mg/L
r
3AS
^
X 70% C
5
/•'
». Anoxic
^ Aerobic I
Is * I
f y o"
^ Aerobic *
^ Anoxic —
,1
|2
J
%
^
^
/
Grit
Chamber
Screens
30% Q
Annerobit
Zone
EQ
Q
The Modified BNR WWTP Operated at Bowie, Maryland, 1988-2008
Figure 6-18. VT2 process schematic.
6.4.1.6 Sequencing Batch Reactor (SBR)
SBRs can be operated to accomplish both biological phosphorus and nitrogen removal by
progressing through anoxic/anaerobic/aerobic phases. Because of the fill-and-draw nature of SBRs, it is
necessary to remove the nitrates remaining from the previous cycle before anaerobic conditions can be
established; thus, the typical treatment progression becomes anoxic/anaerobic/aerobic. Additional
carbon may be needed. See Section 6.2.1.4 for additional discussion of SBRs.
6.4.2 Hybrid Chemical / Biological
Configurations that use some form of biological treatment for nitrogen removal and possibly
phosphorus removal combined with chemical precipitation and advanced solids removal for phosphorus
are popular and typically reliable treatment options for achieving low effluent concentrations. Many
different combinations of the technologies discussed previously can be used. This section presents two
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examples: The Blue Plains Process and the Biological-Chemical Phosphorus and Nitrogen Removal
(BCFS) Process.
6.4.2.1 Blue Plains Process
The Blue Plains process was a retrofit to the existing nitrification activated-sludge process at the
Washington, DC, facility. A new anoxic zone with an HRT of 0.8 hours was created inside the aeration
tank from the nominal 3.3 hours HRT of the total basin (see Figure 6-19). The design sludge age was 13
days. The existing return activated-sludge system remained unchanged in this retrofit. Methanol is fed
directly into this new anoxic zone for a target nitrogen concentration of 7.5 mg/L (Kang et al. 1992;
Sadick et al. 1998). Phosphorus is removed by ferric chloride addition and tertiary filtration (USEPA
2008b).
The effluent concentrations (post tertiary filter) found in the literature for the Blue Plains
process were 7.5 mg/L for TN and 0.12 mg/L for TP on a monthly average basis (USEPA 2008a),
However, that was when only half of the plant was modified for nitrogen removal. When the entire
plant was modified, the effluent TN was less than 5 mg/L (monthly average).
Effluent from
First Stage-
High Rate
Activated Sludse
Aerobic Zone
Nitrification
Anoxic Zone
Denitrification
Aerobic Zone
Nitrification
FeCl, addition
Secondary
Clarifier
Effluent
to
Tertiaiy
Fitter
Methanol Feed
RAS
WAS
Figure 6-19. The Blue Plains Process.
Source: USEPA (2008b), Figure 2-29
6.4.2.2 Biological-Chemical Phosphorus and Nitrogen Removal (BCFS) Process
The Biological-Chemical Phosphorus and Nitrogen Removal (BCFS) process was developed for
the WWTP in Holton, the Netherlands, to achieve low TN and TP effluent concentrations at a relatively
low BOD:TP and BOD:TN influent ratio. The design is similar to the modified UCT process. A sludge
stream is removed from the anaerobic zone for thickening, and ferric chloride is added to the sludge
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thickener to remove phosphate. This provides an advantage over chemical addition to the secondary
clarifier because it does not require the chemical sludge to be recycled. Additional information can be
found on the website of the Dutch Foundation for Applied Water Research (STOWA),
http://www.stowa-selectedtechnologies.nl/ (see Fact Sheets).
6.5 Effluent Filtration
Many varieties of filters can be used in WWTPs. Media range from mono-media such as sand to
multimedia such as sand with anthracite and high density gravel. Other media such as cloth or
membranes (operated under pressure) can be used to capture solids. Filters can be operated in upflow
or downflow mode and can be backwashed using a variety of techniques. Design of multiple filters in
series or filters preceded by a tertiary clarifier can improve solids (and thereby total phosphorus)
removal. Table 6-5 lists some major filtration technologies. A description of each follows the table.
Table 6-5. Matrix of Tertiary Filtration Technologies
Configuration
Tertiary Filtration
Technology
Gravity (down-flow) Filters
Conventional
Deep Bed
Continuous Backwash Upflow Sand Filters
Pulsed Bed Filters
Travelling Bridge Filters
Discfilters (including cloth media)
Membrane Filtration
Section Reference
6.5.1
6.5.2
6.5.3
6.5.4
6.5.5
6.5.6
6.5.1 Conventional Down-flow Filters
These filters consist of fixed-media beds that are typically up to 3 feet in depth and are similar to
filters used to treat drinking water. Media can be single media, dual media, or multi-media. Single
media is typically sand or anthracite. Dual media combines anthracite and sand. Multi-media filters
include a layer of garnet or ilmenite. Differences in specific gravity make it possible to have the largest
particles (anthracite) in the top layer of the filter and the smaller ones (garnet) in the bottom layer.
Flow in these filters is by gravity from the top down. Most of the removal occurs in the top few inches
of the media, but multi-media filters use more of the bed depth for active filtration than mono- or dual
media filters. The filter must be taken off-line periodically to backwash it to prevent clogging and
excessive pressure loss.
Deep bed filters are similar to conventional down-flow filters but have deeper beds and larger
media size. Therefore they have the advantage of longer run times between backwashes. The size of
the media is limited by the ability to backwash the filter. Because these filters are more difficult to
backwash, air scour is necessary to fully clean the filter bed.
6.5.2 Continuous Backwashing Upflow Sand Filters
During operation of the continuous backwashing upflow filter, water is introduced through
risers at the bottom of a deep sand bed. Water flows upward through the sand bed and over an
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overflow weir. Sand and trapped solids flow downward through the filter and are drawn into the
suction of an airlift pipe in the center of the filter. As the sand travels up the airlift pipe, the air scours
the particles and separates the sand from filtered solids. At the top of the airlift pipe, the clean sand
settles back onto the top of the filter and the solids are carried away into a reject line. See Figures 6.8
and 6.9 in Section 6.2.3.1 for cut away diagrams of a continuous backwashing upflow filter.
These filters have the advantages of no moving parts other than the air compressor and less
energy and maintenance requirements than traditionally backwashed filters. They are sometimes
referred to by the trade name Dynasand.
A modification of the standard continuous backwashing upflow filter media is use of sand
coated with hydrous ferric oxide. Ferric oxide coated media provides adsorption in addition to filtration.
One example is the Blue PRO™ system where ferric salt is added prior to the filter to aid in coagulation
and to replace the ferric coating, which is abraded from the sand. Turbulence from the compressed air
knocks accumulated iron and phosphorus along with any solids off the particle as the sand travels
upward. The iron, phosphorus, and particles are wasted, while the clean sand is deposited on the top of
the bed. Benish et al. (2006) report that "an average effluent TP of 0.06 mg/L was observed during a
pilot study in Moscow, Idaho, and near 0.02 mg/L was achieved in a pilot study in Coeur d'Alene, Idaho."
EPA Region 10 (2007) summarized results of demonstration of the Blue PRO™ system at a 0.25 MGD
Hayden Wastewater Research Facility. Secondary effluent with a total phosphorus concentration of
between 1 and 4 mg/L was treated through a two-single-pass Blue PRO™ filtration system. The average
effluent concentration for the entire test period (2005 to 2007) was 0.014 mg/L.
6.5.3 Pulsed Bed Filters
Pulsed bed filters are shallow filters with an unstratified fine sand media. An air pulse disturbs
the media and allows penetration of solids into the media bed, allowing the entire filter bed to be used
for removal of solids. The pulse is designed to expand the filter operation and reduce the number of
backwash cycles, although the filter must still be periodically backwashed to remove the solids.
6.5.4 Traveling-Bridge Filters
Traveling-bridge filters consist of long shallow beds of granular media divided into cells.
Wastewater is applied to the top of the media and flows downward. Each cell is individually
backwashed by a traveling bridge while the other cells continue to operate. The bridge uses filtered
water to backwash the filters. This type of filter is used for the removal of effluent TSS, which can
enhance nutrient removal because TSS contain both nitrogen and phosphorus. Effluent TSS is usually
between 5 and 10 mg/L, which is not suitable for enhancing phosphorus removal below 0.5 mg/L.
6.5.5 Discfilters
Discfilters are a series of parallel mounted disks used to support a cloth filter media. Water
enters a central tube and flows out between the two layers of cloth in each disk. The disks rotate and
are normally 60 to 70 percent submerged. The portion above the water is backwashed using spray
nozzles.
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The cloth media disk filter is similar to the discfilter described above. In this case, the water
flows from the outside of the submerged cloth disks and into a center pipe. Disks continue to rotate
during backwash, as water is pulled through the media from the disc using suction heads.
Discfilters and cloth media disk filters have limited solids loading capacities and are thus most
often used for BNR applications rather than following chemical addition for phosphorus removal.
6.5.6 Membrane Filters
Membrane systems use a pressure head to drive water through a permeable membrane.
Membrane filters are typically classified by their pore size, which in turn determines the size of the
particles they exclude and guides the pressure requirements. Microfiltration, ultrafiltration,
nanofiltration, and reverse osmosis (RO) membranes remove increasingly smaller particles. Biological
treatment that includes microfiltration and ultrafiltration typically remove 3 to 6 logs of bacteria and 95
percent or more BOD, along with most particles (WEF 2006). Nanofiltration removes nearly all particles,
including some viruses. RO removes all particles as well as most large dissolved constituents, and is
capable of producing pure water. Typically, the water must be pre-treated before using RO membranes.
Pre-treatment could be conventional filters, cartridge filters, or larger pore membrane filters. The
energy cost for applying the pressure head and the need to replace membranes make membrane
filtration a more expensive technology. It can achieve very low concentrations of nutrients and other
contaminants, however, and is common in water re-use projects.
Membranes can be configured a number of ways including hollow fiber, spiral wound, plate and
frame, cartridge, or in pressure vessels. Membranes can foul from organics, biological growth, or metals
in the wastewater. Disinfection may be required to prevent biological fouling.
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6.6 Sidestream Management
Solids handling processes result in a liquid sidestream that is generally recycled to the head of
the treatment plant and combined with the influent wastewater. The concentration of nitrogen and
phosphorus in the sidestream depends on the influent concentrations and the main stream removal
technology. It also depends highly on the solids treatment technology and equipment, with the highest
nutrient concentration resulting from anaerobically digested or thermally conditioned sludge (WEF and
ASCE 2006). As will be discussed in Chapter 10, characterizing sidestreams (e.g., nutrient loadings and
when they are discharged) is very important when designing for nutrient removal, particularly for plants
wanting to consistently achieve low effluent concentrations.
Equalization of sidestreams is a proven management technique to reduce the impact of the
sidestreams on the main treatment train. An equalization tank can be sized to hold the total volume of
recycle flows for a certain period of time so that they can be sent to the treatment system at selected
rates or times to minimize the impact on the treatment process (e.g., at night when total plant flows and
nutrient loadings are low). An equalization tank can also be designed to settle and remove solids to
reduce the loads back to the plant, but the settled solids have to be collected and processed. Aeration
within the tank or as a separate process has also been shown to reduce the ammonia load back to the
plant by up to 50 percent given that detention time is sufficient and the tank can be fully mixed (WEF
and ASCE 2006).
Over the last several years, patented processes have been designed to remove nitrogen from
recycle streams. The following examples are discussed in this section, but note that this is not a
complete list and new processes are continually being developed:
• Bioaumentation (e.g., InNitri® or BABE®)
• BAR
• MAUREEN
• SHARON®
• ANAMMOX®
• DEMON®
• CANON
• OLAND
InNitri® and BABE® use bio-augmentation (i.e., the seeding of nitrifiers) to enhance nutrient
removal. InNitri® (Inexpensive Nitrification) grows nitrifiers in a small, sidestream aeration-clarifier
system by treating high ammonia, high temperature water from digestion and dewatering to obtain
complete nitrification of the ammonia. The waste activated sludge (WAS) from the sidestream system is
discharged to the main aeration tank to supplement the nitrifiers in that reactor, thereby reducing the
SRT needed to maintain nitrification at low temperatures. InNitri® does not remove nitrogen but
reduces the volume needed in the main stream reactor to accomplish nitrification.
The BABE® (BioAugmentation Batch Enhanced) process uses a single batch reactor to treat a
mixture of high ammonia sidestreams from dewatering, RAS from the main biological treatment system,
and residual biomass remaining in the reactor from the preceding cycle. The high ammonia
concentration is completely nitrified and then partially denitrified during each cycle of the batch reactor
to obtain removal of nitrogen. The RAS is partially seeded with nitrifiers and denitrifiers by the BABE
Nutrient Control Design Manual 6-32 August 2010
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process during this operation. The BABE process is designed to operate between 20 and 25° C, and
temperatures below 20° C will dramatically increase the reactor volume needed.
The Bio-Augmentation Regeneration/Reaeration (BAR) process was developed in the United
States and is identical to the Regeneration-DeNitrification (R-DN) process developed independently in
the Czech Republic. It works by recycling ammonia-laden filtrate or centrate from dewatering of
aerobically digested sludge to the head of the aeration tank. The sidestream is fully nitrified which
seeds the aeration tank with additional nitrifying bacteria and allows for reduced SRT. There are
numerous full scale applications in the Czech Republic, United States, and Canada.
The Mainstream AUtotrophic Recycle Enabling Enhanced N-removal (MAUREEN) Process was
developed for the two-sludge treatment configuration at the Blue Plains Advanced Wastewater
Treatment Plant in Washington, DC. The process involves recycling of WAS from the nitrification stage
to a reactor where the AOBs are used to oxidize the centrate ammonia to nitrite. Then the nitrite is
reduced to nitrogen gas, and the AOBs are sent to the high-rate carbon removal stage for
bioaugmentation of the first sludge stage activated sludge to accomplish some nitrification in that stage.
The SHARON® (Single reactor for High Activity Ammonia Removal Over Nitrite) process takes
place in a single reactor operated without biomass recycle, i.e., the SRT is equal to the HRT. Developed
to treat high ammonia concentration sidestreams from sludge digestion, the system is designed to
convert ammonia to nitrite, then denitrify the nitrite to nitrogen gas, thus avoiding the conversion of
nitrite to nitrate and the need to reduce nitrate back to nitrite to obtain nitrogen removal. This process
is accomplished by operating the reactor at high temperatures and selecting an HRT (equals SRT) that
will wash out the slow-growing nitrite oxidizers. The process is typically operated at 30 to 40° C with an
HRT range of 1 to 2 days. By preventing the conversion of nitrite to nitrate, the SHARON® process
reduces the amount of oxygen required for nitrification by 25 percent and then reduces the amount of
COD needed for denitrification by 40 percent, compared to full nitrification and denitrification. See
Chapter 4, Section 4.5 for additional discussion on the principles of the SHARON® process.
The ANAMMOX® (ANaerobic AMMonia OXidation) process uses only autotrophic bacteria to
accomplish nitrogen removal, i.e. denitrification as well as nitrification, and, because ammonium is used
as an electron donor, COD does not have to be added to accomplish nitrogen removal. In the
conventional ANAMMOX process the first stage is similar to the SHARON® process except that only 50
percent of the ammonia is oxidized to nitrite, and a 50-50 mixture of ammonium and nitrite is sent to
the second reactor, where the anammox bacteria use ammonium as an electron donor and reduce
nitrite to nitrogen gas. Thus, the oxygen requirement is only 50 percent of that required for complete
oxidation of ammonia to nitrite by the SHARON® process, or only 37.5 percent of the amount needed
for nitrification to nitrate, assuming nitrification of the entire ammonium load. In addition, the need for
organic carbon is completely eliminated. It has been reported that this process reduces the operating
costs by 90 percent compared to standard nitrification-denitrification. See Chapter 4, Section 4.12 for
additional discussion on the principles of the ANAMMOX® process.
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The combination of partial nitrification and Anammox processes can also be carried out in a
single reactor and has been given different names: DEMON (DE-amMONification), OLAND (Oxygen-
Limited Autotrophic Nitrification-Denitrification) or CANON (Completely Autotrophic Nitrogen removal
over Nitrite process). Under oxygen-limited conditions a co-culture of aerobic and anaerobic
ammonium-oxidizing bacteria can be maintained as a co-culture in a single reactor.
The DEMONR process uses pH and DO control in an SBR to maintain both nitrification to NO2-N
and nitrite reduction by anammox bacteria to convert ammonia to nitrogen gas. It allows both of these
types of bacteria to co-exist in one single reactor volume, despite the completely different growth
characteristics of these bacteria. Within the DEMON process, approximately 50 percent of the ammonia
is nitrified to nitrite. External carbon addition is typically not needed.
Another variation of the above application of the anammox process in a single SBR tank is the
CANON process (Vazquez-Padin et al., 2009). In this application, air is pulsed to the reactor at a rate and
frequency to maintain low microaerophilic DO concentrations.
The OLAND process had been used in laboratory studies with SBR and fixed film applications
(Pynaert et al., 2002, Clippelier at al., 2009). For SBR applications low DO and low volumetric removal
rates were considered critical to maintain an anammox culture in a granular floe. The fixed film
application showed that the anammox reaction could be maintained in a rotating biological contactor
with a liquid DO concentration of near 1.0 mg/L.
6.7 Technology Performance
For a given biological nutrient removal technology, the daily composite effluent nitrogen and
phosphorus concentrations will vary from day to day and vary among sites due to a number of factors.
Commonly recognized factors that affect BNR plant performance are summarized below.
• Wastewater characteristics are of particular importance in BNR processes. The characteristics
and relative amount of influent biodegradable COD are key to biological denitrification for
nitrogen removal and also for the performance and capacity of enhanced biological phosphorus
removal. This and other important wastewater characteristics are given as follows:
- Influent rbCOD:TP and BOD:TN ratios
Minimal temperature and temperature variations
Daily and seasonal variations in influent flowrates
Daily and hourly variations in flow and consitutent concentrations
Seasonal variations in flow and constitutent concentations
Industrial contributions (e.g., toxins)
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• BNR Treatment plant design and configuration involve a number of complex process
considerations and equipment selection. For example, choosing a high SRT may be beneficial for
some processes such as nitrification, but not for enhanced biological phosphorus removal. It
should be noted that full-scale plant data used to judge BNR treatment performance is typically
obtained from plants that are to varying degrees below their design capacity. Key factors that
can affect performance are as follows:
Design conservatism in SRT and HRT
- Treatment goals (related to permit level and averaging period)
Process flexibility
Solids processing methods
- Type and methods of handling recycle flows
Equalization capability
Sidestream treatment of recycle flows
• Installed equipment
Equipment performance and reliability
Equipment maintenance requirements
Redundancy
• Operation
- Staffing
Instrumentation and process control methods
When evaluating effluent concentrations for a specific BNR technology, it is necessary to
recognize that the effluent performance varies daily as a function of influent wastewater, process
conditions, operating methods, and biological treatment variations. Thus,the effuent is more aptly
characterized in terms of a statistical performance.
Neethling et al. (2009) introduced a method for using a statistical approach to describe process
performance. In this approach, the treatment plant or technology performance is tied to the statistical
rank to express the probability of achieving a certain performance. Building on this statistical approach,
the term Technology Performance Statistic (TPS) was used at a WEFTEC workshop (WEF/WERF, 2009) to
assess the performance of full scale treatment plants.
The Technology Performance Statistics (TPS) describes the performance of a technology or
process or plant under specific conditions. The TPS is determined from performance data and is linked
to the operational conditions during which the data were collected (pilot, full scale, summer, winter,
excess capacity available, SRT, etc). The conditions must also include external factors that impact the
technology, industrial loadings, seasonality, absence of recycle streams, etc. In addition, the TPS
established using past performance, is tied to the treatment objectives or permit limits.
As will be described in more detail in Chapter 7, permit limits for nitrogen and phosphorus can
be based on annual average values, quarterly averages, seasonal averages, maximum monthly averages,
maximum weekly values, or at their most stringent, maximum daily recorded levels. Limits may be
concentration or mass based. Many permits include multiple types of limits for the effluent parameters.
Nutrient Control Design Manual 6-35 August 2010
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Annual average permit limits for TN and TP clearly provide the greatest flexibility in operation
and the least risk of violations. In temperate regions, the BNR system can be operated with relatively
high effluent nitrogen concentrations during the winter months, and then operated to obtain very low
concentrations during the late summer-early fall months. Seasonal averages provide a similar but lesser
amount of flexibility. Monthly or weekly limits necessitate that effluent limits be met on a consistent
basis regardless of seasonal operating conditions. The shorter the time period the less forgiving the limit
is for periodic upsets of the biological system, such as partial nitrifier washout during storm events. The
end result is that WWTPs are often operated to produce effluent concentrations lower than the
required value, providing a margin of safety and protecting against a single day exceedance and
potential permit violation. This approach may require higher operating costs or more expensive
technologies than would be designed and constructed for the longer term averages.
The remainder of this section provides a generalized discussion of the effectiveness of nutrient
removal technologies followed by a description of recent research to identify reliable effluent
concentrations using data from full-scale treatment plants.
6.7.1 Removal Efficiencies of BNR Technologies - General Discussion
Nitrogen Removal
For biological nitrogen removal, the attainable effluent limit for a suspended growth process
depends on the number of anoxic zones and the nitrate recycle ratio to the influent flow. The basic
configuration of a single anoxic and aerobic zone (the MLE system), has historically achieved 70 to 90
percent TN removal. TN removal by this basic configuration has been improved by the use of attached
growth filters for nitrification (e.g, BAF) or attached growth media in the anoxic and/or aerobic basin to
increase SRT and reduce the risks of washout. A second anoxic zone after the aerobic zone (e.g., the 4-
stage Bardenpho process) can achieve additional TN removal by denitrifying that portion of the flow that
is not recycled to the first anoxic zone. Because the BOD has been removed in the aerobic zone,
denitrification in a post-anoxic zone often requires a supplemental carbon source especially in colder
regions where the endogenous denitrification rate is low.
Nitrification processes can be optimized by controlling influent ammonia loading (from
wastewater or recycle streams) and flow to prevent washout of nitrifiers. Configuration of swing zones
that can be operated in anoxic or aerobic modes can help ensure consistent removal under varying
operating conditions (e.g., seasonal variations in influent loading, temperature changes, flow variation).
Similarly, online monitoring and process control of aeration systems and recycle streams can allow for
more consistent performance of systems. Sidestream treatment such as bioaugmentation can be used
to seed the activated sludge process with an external source of nitrifiers to obtain complete nitrification
at lower SRTs.
For removal of total nitrogen to very low levels (3 mg/L or less), technologies may be limited by
biologically resistant dissolved organic nitrogen (rDON), which is typically between 0.5 and 1.5 mg/L.
Research is ongoing into the design and operating conditions that can be used to minimize rDON.
Process optimization and automated control systems become very important in achieving these low TN
levels. Denitrification filters with supplemental carbon (e.g., methanol) addition can provide enhanced
TN removal consistently to low effluent levels.
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Phosphorus Removal
For biological phosphorus removal, the attainable effluent concentration is fundamentally
determined by the rbCOD:TP ratio entering the anaerobic zone, minus the mass of electron acceptors
(i.e., DO and nitrates) entering the anaerobic zone. If the rbCOD:TP ratio is favorable ( 16:1 or higher)
and the recycle of nitrate and DO to the anaerobic zone is carefully controlled, low effluent
concentrations can be reliably obtained using almost any of the BPR technologies. If VFAs are added
from sludge fermentation, lower rbCOD:TP ratios can achieve the same results, with the variability
depending upon the quantity and mix of VFAs formed by the fermentation. Unsettled sewage will
produce lower effluent soluble phosphorus concentrations than settled sewage because settling
changes the COD to TP ratio of the wastewater in an unfavorable direction.
Phosphorus removal by chemical addition (with or without biological phosphorus removal) can
reliably achieve low effluent TP levels. Metal salts such as alum and ferric chloride are commonly used.
Complete mixing of the chemical at the dosing point and adequate flocculation assure that the metal
reacts with phosphate to form a precipitant. The key factor to achieving low levels is the method of
liquid/solids separation. Tertiary filtration, in particular deep bed, dual media systems and membranes,
provide nearly complete removal of particulate matter and can achieve very low effluent TP levels.
Advanced processes such as ballasted high-rate clarification (BHRC) and filtration using iron-oxide
coated media have been successful at removing TP to low levels in a number of full-scale applications.
For reliably removing phosphorus to very low levels, the recalcitrant fraction (i.e., unreactive
fraction that goes through the treatment process unchanged) may become limiting depending on the
nature of the influent wastewater. In addition, variability in measurement of phosphate concentrations
less than 0.02 mg/L could be a limiting factor. See Chapter 3, section 3.7 for additional discussion.
6.7.2 Technology Performance Statistics based on Full-Scale Operating Data
The Water Environment Research Foundation (WERF) sponsored a project to evaluate the
performance capabilities of nutrient removal processes by collecting data from some of the best
performing nutrient removal WWTPs in the county to improve understanding of achievable effluent
levels for nutrients and key factors affecting plant performance to these levels (WERF project no.
NUTR1R06J, 2009). Preliminary results of the project were presented at the 2009 Annual Water
Environment Federation Technical Exhibition and Conference (WEFTEC) (Neethling, et al. 2009; Parker,
et al. 2009; Bott, et al. 2009).
The facility evaluations were based on complete operating data and analytical information
provided by plant managers covering a 3 year period. No special sampling or operational changes were
made for the performance evaluations. Statistical methods were used to determine the probability
statistics for nitrogen and/or phosphorus removal performance at each plant. Probability plots were
developed using normal and log-normal distributions. See Bott et al. (2009) for a detailed discussion of
statistical analyses of plant data.
The WERF researchers recognized the variability in nutrient removal performance that exists in
all plants due to variations in wastewater characteristics and operating conditions, and consequently it
was not considered practical to describe the achievable limits for a technology based on a single
numerical value. Thus, the achievable performance was described as Technology Performance
Statistics (TPS). In this way the performance levels could be defined in terms of frequency and
Nutrient Control Design Manual 6-37 August 2010
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reliability. In addition, IPS results should consider the conditions under which the data were collected
such as process configurations, treatment objectives, existing versus design loadings, solids handling
processes, etc. See (Neethling, et al. 2009) for more information.
Three IPS levels can be used to define the best, median, and reliably achievable performance:
• Best performance: The lowest concentration achieved by the plant during a 14-day period (TPS-
14d). Neethling, et al. (2009) provides the rationale for this value, explaining that a 2-week
period would capture two weekly cycles, eliminate outlying data, identify a high level of
treatment that can occasionally be achieved, and perhaps also capture a cycle of at least one
sludge age in a typical plant operation. The TPS-14d can be calculated in two ways: (1)
calculating a running 14-day average and determining the minimum value, or (2) calculating the
3.84th percentile.
• Median performance: The concentration achieved on a statistical median basis (TPS-50%)
approximating average performance. This value indicates the median performance of the
technology that can be achieved on an annual basis with no margin of safety. It is statistically
exceeded every other year and requires a margin of safety to provide reliable performance to
comply with discharge limits.
• Reliably Achievable Performance: May be the 90th, 95th, or 99th percentile depending on the
permit averaging period. Neethling, et al. (2009) notes that the 91.7th percentile (11/12
percent) translates to a plant with monthly limits exceeded one month per year on a statistical
basis. The TPS-95% indicates a monthly value that is exceeded 3 months in a 5 year cycle.
Figure 6-20 illustrates technology performance statistics of a phosphorus removal plant. The
data from this facility follows a good approximation of the typical log-normal distribution. When
determining performance statistics, the rank is used to calculate performance to account for the actual
performance distribution as reported by the technology. The TPS-3.84% or best performance for this
dataset is 0.040 mg/L This value excludes the best reported value of 0.002 mg/L which is likely an
outlier. The 50th percentile performance statistic of 0.08 mg/L reflects the median value and the 95th
percentile statistic of 0.23 mg/L the reliable performance. The 95th percentile effluent P concentration
is 3 times the median performance and 6 times the best performance.
The distribution in Figure 6-20 illustrates the importance of the averaging period in determining
plant performance. This example plant demonstrates that it can achieve 0.08 mg/L on average, yet the
reliable monthly value is three times higher at 0.23 mg/L. The longer averaging period increases the
ability of the plant to achieve the limit. A shorter averaging period requires a higher safety factor to
increase the ability to achieve the limit reliably.
Nutrient Control Design Manual 6-38 August 2010
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10.00
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0.1% 1% 10% 25% 50% 75% 90% 99% 99.9%
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Figure 6-20. Probability plot of secondary effluent phosphorus data showing 3.84th, 50th, and 95th percentiles of the data
showing a log-normal distribution line and determining the percentiles based on this log-normal distribution (Example data).
Summary of Performance Data
IPS values for the facilities participating in the WERF study are summarized in Figure 6-21 for
nitrogen removal and Figure 6-22 for phosphorus removal. The figures show the best, median, and
reliable (95%) IPS values.
Nutrient Control Design Manual
6-39
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Nutrient Control Design Manual
6-40
/August 2010
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Nitrogen removal facilities are grouped into three categories:
• Separate Stage processes. These nitrogen removal plants use separate unit processes for
nitrification and denitrification. A carbon source such as methanol is added to the denitrification
stage.
• Combined nitrogen removal processes represent conventional, multiple zone activated sludge
processes designed for biological nitrogen removal. These processes include MLE, Bardenpho,
step feed, or other processes. Effluent filters are used in these plants without added carbon -
the filter removes particulate nitrogen but does not provide denitrification.
• Multiple Stage processes are those that use conventional biological nitrogen removal (as in the
combined processes) with additional nitrogen removal in a denitrification filter with carbon
addition.
The performance of the facilities in Figure 6-21 show the three TPS values in groups. The graphs show
that combined processes are not achieving as low effluent nitrogen values as separate and multistage
processes. In addition, the variability of the combined (or conventional) processes is high: for two of the
facilities, the reliable TN concentration is about 3 times the median performance. Multistage and
separate stage facilities achieve lower effluent TN concentrations and show less variability with reliable
TPS-95% values about 2 times the average. The reason for the improved performance is in the ability to
add external carbon to overcome excursions in performance and maintain the performance levels at the
target by adjusting the carbon addition.
Phosphorus removal facilities are grouped into four categories depending on the main removal
process (biological or chemical) and depending on addition of tertiary chemical addition:
• IB represents Biological Phosphorus Removal with a polishing filter, but without chemical
addition to the filters.
• 1C represents Single Chemical Phosphorus Removal in the primary or secondary process with a
polishing filter but without chemical addition to the filters.
• 2B represents Biological Phosphorus Removal with polishing filter with chemical for phosphorus
removal.
• 2C represents Single Chemical Phosphorus Removal in the primary or secondary process with a
polishing filter with chemical for phosphorus removal.
The performance of the facilities in Figure 6-22 show the three TPS values in groups. The graph shows
that 2 stage processes are achieving lower effluent phosphorus values than single stage processes.
There appears to be little difference in the performance of plants using biological and chemical removal
for the primary and secondary process. The variability in the phosphorus removal is significant. The
reliable performance is 1.5 to 4 times higher than the average.
Nutrient Control Design Manual 6-41 August 2010
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6.8 Factors for Simultaneously Achieving Low Nitrogen and Phosphorus Effluent
Concentrations
Some unique factors apply when plants are attempting to simultaneously achieve low nitrogen
and phosphorus in the plant effluent. Key factors are discussed below. For additional discussion, see
Section 2.4.1 of the Municipal Nutrient Removal Technologies Reference Document (USEPA 2008b).
The microorganisms responsible for biological removal of phosphorus and nitrogen compete for
bio-available carbon in the waste stream. The available carbon in the influent or return side streams can
be used for both biological phosphorus removal and denitrification, but is likely to be insufficient to
simultaneously achieve very low effluent levels of both, for instance, below 4.0 mg/LTN and 0.5 mg/L
TP. This limitation can be overcome by adding a supplemental organic carbon source such as methanol
or some other alternative to the second anoxic zone or VFAs to the either the anaerobic or second
anoxic zones, or both. Plants that are required to meet very low effluent phosphorus limits will most
likely require both the addition of an alternative carbon source to the anaerobic zone and tertiary
filtration.
Phosphorus is a necessary nutrient for microbial growth. If TP is reduced too low before a
biological process like nitrification or denitrification, the growth of the microorganisms could be
inhibited, necessitating the addition of supplemental phosphorus. This is most likely to occur when
tertiary denitrification filters are used; however, inhibition of nitrification also has been observed in
plants practicing chemical phosphorus removal in the primary clarifiers.
In plants with a single anoxic zone for nitrification, nitrates may be returned in the RAS stream
to the anaerobic zone. This can reduce BPR because some heterotrophic microorganisms use nitrates to
grow and will consume organic substrate instead of the PAO. To address this issue, the following design
options can be used:
• Convert the last part of the aeration zone into a deoxygenation section by reducing the aeration.
Care must be taken not to cause rising sludge in the secondary clarifier.
• Divert the RAS to a small tank before returning it to the anaerobic zone, such as in the RON
configuration. The tank would reduce the DO and nitrate in the RAS, which would further
improve the performance of the anaerobic zone. The retention time, however, would have to be
limited to prevent secondary release of phosphorus.
• Divert the RAS to the anoxic zone rather than to the anaerobic zone, where it will both
deoxygenate and denitrify, and then recycle to the anaerobic zone, as in the UCT/VIP
configurations.
Plants should be designed and operated to avoid secondary release. Anaerobic or anoxic zones
that are too big remove organic carbon and nitrates in less time than the HRT. This can result in the
release of phosphate without poly-(3-hydroxy-alkanoate (PHA) storage, which means that there will be
insufficient stored substrate to take up the released phosphorus in the aerobic zone. Plant operators
can consider adding a chemical to primary sedimentation tanks (e.g., ferric chloride) in the winter to
remove more solids and reduce the needed SRT to achieve the same nitrification levels.
Nutrient Control Design Manual 6-42 August 2010
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If nitrate removal has been enhanced in the secondary process, phosphorus can be released
from the sludge in the secondary clarifier if DO depletion occurs. This might not be observed
immediately but could result in a buildup of phosphorus in the secondary treatment system that would
eventually raise the effluent TP levels. If this is a possible issue, plant operators should consider
biological nitrification and phosphorus removal in a tertiary process using chemical addition.
Temperature affects biological nitrogen and phosphorus removal processes similarly although
nitrification is more sensitive than BPR. In colder temperatures, fermentation might not provide
sufficient VFAs for the BPR process. Sidestream fermentation has been used very successfully to
augment VFAs in the anaerobic zone during winter operation. The rate of nitrification decreases
significantly with decreasing temperatures, with rates 50 percent lower for every 8 to 10°C drop in
temperature (WEF and ASCE 2006). To maintain nitrification of ammonia (which is the key first step to
nitrogen removal), plants can design swing zones and aerate those zones during periods of low
temperatures. Denitrification rates also decrease with decreasing temperature, although not as rapidly.
Denitrification rates at low temperatures should also be considered when designing and operating
nutrient removal systems.
Both nitrification and chemical phosphorus removal consume alkalinity. Low alkalinity can
inhibit nitrification because it is the carbon source for the nitrifiers, and also because low alkalinity
usually causes low pH, which can inhibit nitrifiers. Therefore, even when nitrogen removal is not
required but phosphorus removal is, practicing denitrification is usually the most economical approach
because denitrification restores 50 percent of the alkalinity consumed during nitrification. Of course,
this is a useful strategy only if nitrification is already being practiced. Note that chemical precipitation
will not be complete if sufficient alkalinity is not present.
6.9 References
Barnard, J.L 1984. Activated Primary Tanks for Phosphate Removal. Water SA. 10(3): 121-126.
Barnard, J.L 2006. Biological Nutrient Removal: Where We Have Been, Where We are Going? In
WEFTEC 2006.
Barnard, J.L, H. Phillips, B. Sabherwal, C. deBarbadillo. 2008. Driving Membrane Bio-Reactors to Limit
of Technology. In WEFTEC 2008.
Benish, M., D. Clark, A.Z. Gu, J.B. Neethling. 2006. How Low Can You Go? Achieving Extremely Low
Effluent Phosphorus in Wastewater Treatment. Water and Wastes Digest, October 2006. Volume 46
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Bott, C.B., S. N. Murthy, T. T. Spano, and C.W. Randall. 2007. WERF Workshop on Nutrient Removal:
How Low Can We Go and What is Stopping Us from Going Lower? Alexandria, VA: WERF.
Bott, C.B, D. Parker, J. B. Neethling, A. Pramanik, and S. Murthy. 2009. WEF/WERF Cooperative Study of
BNR Plants Approaching the Limit of Technology: II. Statistical Evaluation of Process Reliability.
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Clark, D.L, Hunt, G., Kasch, M.S., Lemonds, P.J., Moen, G.M., Neethling, J.B. (2010) Nutrient
Management Approaches to Protect Water Quality. Volume 1 Review of Existing Regulatory Practices,
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Clippelier, H.D., S.E. Vlaeminck, M. Carballa, and W. Verstraete, 2009. A Low Volumetric Exchange Ratio
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100: 5010-5015.
DeBarbadillo, C, J. Barnard, S. Tarallo, and M. Steichen. 2008. Got Carbon? Widespread biological
nutrient removal is increasing the demand for supplemental sources. Water Environment &
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Dold, P., I. Takacs, Y. Mokhayeri, A. Nichols, J. Hinojosa, R. Riffat, C. Bott, W. Bailey, and S. Murthy.
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80(5): 417-427. WEF.
Ellis, E.P. and A.M. Cathcart. 2008. Selection, Installation, Startup and Testing of the World's First Full-
Scale CoMag Phosphorus Reduction Tertiary Treatment System. Presented at WEFTEC 2008. Chicago,
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EPA Region 10. 2007. Advanced Wastewater Treatment to Achieve Low Concentration of Phosphorus.
EPA Region 10. EPA 910-R-07-002.
Feldthusen, M. 2004. Continuous Sand Filters: Tertiary WWT and Other Applications. SAWEA-
Workshop. March 22, 2004.
Holloway, R., H. Zhao, T. Rinne, G. Thesing, J. Parker, and M. Beals. 2008. The Impact of Temperature
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Hu, Z. 2001. Eternal Nitrification Biological Nutrient Removal Activated Sludge Systems: Development
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Rondebosch, 7700, South Africa.
Kang, S.J., W.F. Bailey, and D. Jenkins. 1992. Biological Nutrient Removal at the Blue Plains Wastewater
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Neethling, J.B., B. Bakke, M. Benisch, A. Gu, H. Stephens, H.D. Stensel, and R. Moore. 2005. Factors
Influencing the Reliability of Enhanced Biological Phosphorus Removal. Alexandria, VA: WERF and IWA
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Neethling, J.B., D. Stensel, D. Parker, C. Bott, S. Murthy, A. Pramanik, and D. Clark. 2009. What is the
Limit of Technology (LOT)? A Rational and Quantitative Approach. Presented at WEFTEC 2009,
Orlando, FL.
Parker, D., C. Bott, J. B. Neethling, A.Pramanik, and S.Murthy. 2009. WEF/WERF Cooperative Study of
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Pynaert, K., R. Sprengers, J. Laenen, and W. Verstraete, 2002. Oxygen-limited Nitrification and
denitrification in a Lab-scale Rotating Biological Contactor. Environment Technology. 23(3): 353-362.
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Pearson, J.R., D.A. Dievert, D.J. Chelton, and M.T. Formica. 2008. Denitrification Takes a BAF: Starting
Up the First Separate Biological Anoxic Filter in Connecticut Requires Some Problem-Solving and
Know-How. Water Environment & Technology. Alexandria, VA: WEF. 20(5): 48-55.
Randall, C.W., J. Barnard and D. Stensel. 1992. Design and Retrofit of Wastewater Treatment Plants for
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Reardon, Roderick D. 2005. Tertiary Clarifier Design Concepts and Considerations. In WEFTEC2005.
Sadick, T., W. Bailey, G. Daigger, and M. McGrath. 1998. Large-Scale Nitrogen Removal Demonstration
at the Blue Plains Wastewater Treatment Plant Using Post- Denitrification with Methanol. Presented
at IAWQ 19th Biennial International Conference on Water Quality. Vancouver, B.C., Canada. June
1998.
Sen, D., C. Randall, and T. Grizzard. 1990. Biological Nitrogen and Phosphorus Removal in Oxidation
Ditch and High Nitrate Recycle Streams. Pub CDP/TRS 47/90. USEPA Chesapeake Bay Program.
Sen, S., V Occiano, P. Wong, and A. Langworthy. 2008. Comparing Implementation of MBBR versus BAF
on a Space Constrained Site. In WEFTEC 2008.
STOWA. 2006. BCFS Process Sheets. Website updated 13 June 2006. Available online:
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Trivedi, Hiren and Nicolas Heinin. 2000. Simultaneous Nitrification / Denitrification by Monitoring
NADH Fluorescence in Activated Sludge. In WEFTEC2000.
USEPA. 2003. Wastewater Technology Fact Sheet: Ballasted Flocculation. Office of Waste Management.
Municipal Technology Branch. EPA 832-F-03-010
USEPA. 2007a. Wastewater Management Fact Sheet: Denitrifying Filters. EPA 832-F-07-014.
USEPA. 2007b. Wastewater Management Fact Sheet: Membrane Bioreactors. Available online:
http://www.epa.gov/owm/mtb/etfs membrane-bioreactors.pdf
USEPA. 2008a. Emerging Technologies for Wastewater Treatment and In-Plant Wet Weather
Management. EPA 832-R-06-006. Available online: http://www.epa.gov/OW-
OWM.html/mtb/emerging technologies.pdf
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Management, Municipal Support Division. EPA832-R-08-006. Available online:
http://www.epa.gov/OWM/mtb/publications.htm
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Reactor Systems. EPA 625-R-00-008. Website updated: 14 April 2008. Available online:
http://www.epa.gov/nrmrl/pubs/625r00008/html/html/tfs3.htm
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Vaxquez, J.R, M. Figueroa, I. Fernandez, A. Mosquera-Corral, J.L Campos, and R. Mendez. 2009. Post-
treatmente of Effluents from Anaerobic Digesters by the Anammox Process, Water Science and
Technology, 60(5): 1135-1143.
Verma, M., S.K. Brar, J.F. Blais, R.D Tyagi, and R.Y. Surampalli. 2006. Aerobic Biofiltration Processes-
Advances in Wastewater Treatment. Pract. Periodical of Haz., Toxic, and Radioactive Waste Mgmt.
10: 264-276.
Vidal, A., R. Combeau, J. Costa, X. LeTallec. 1997. Biostyr Improvements: Control Process of the
Aeration. Annual Report of Anjou Research on Biostyr Process. OTV, France.
WEF and ASCE. 2006. Biological Nutrient Removal (BNR) Operation in Wastewater Treatment Plants -
MOP 29. Water Environment Federation and the American Society of Civil Engineers. Alexandria, VA:
WEFPress.
WEF. 2006. Membrane Systems for Wastewater Treatment. Alexandria, VA: WEFPress.
WEF. 2005. Manual of Practice FD-8: Clarifier Design, 2nd Edition. Alexandria, VA: WEF.
WEF and ASCE. 2010. Design of Municipal Wastewater Treatment Plants - WEF Manual of Practice 8 and
ASCE Manuals and Reports on Engineering Practice No. 76, 5th Ed. Water Environment Federation,
Alexandria, VA, and American Society of Civil Engineers Environment & Water Resources Institute,
Reston, Va.
WERF. 2000a. Technology Assessments: Nitrogen Removal Using Oxidation Ditches. Alexandria, VA,
WERF.
WERF. 2000b. Investigation of Hybrid Systems for Enhanced Nutrient Control. Final Report, Collection
and Treatment. Project 96-CTS-4. Alexandria, VA: WERF.
WE&T. 2007. Plant Profile: Warrenton Wastewater Treatment Plant. Water Environment & Technology.
Alexandria, VA: WEF. Available online:
http://www.wef.org/ScienceTechnologvResources/Publications/WET/07/07Mav/07MavPlantProfile.ht
rn
Young, T., S. Crosswell, and J. Wendle. 2008. Comparison of Nitrogen Removal Performance in SBR
Systems. In WEFTEC2008.
Zhu, S. and S. Chen. 2002. The Impact of Temperature on Nitrification Rate in Fixed Film Biofilters.
Aquacult. Eng. 26,221-237.
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7. Establishing Design Objectives
Chapter 7 covers:
7.1 Introduction
7.2 Characterizing Existing Treatment
7.3 Design Flow Rates
7.4 Wastewater Characteristics
7.5 Target Effluent Concentrations for Nitrogen and Phosphorus
7.6 Goals for Reliability, Sustainability, and Process Flexibility
7.7 Solids Handling Options
7.8 Site Constraints
7.9 Selecting an Overall Process Design Factor
7.10 References
7.1 Introduction
Establishing design criteria is a critical first step in upgrading or retrofitting a wastewater
treatment plant (WWTP) for nutrient removal. Design criteria are the conditions under which the
WWTP must operate following the upgrade or retrofit. They include factors such as projected influent
flow rate and mass loading, effluent requirements (e.g., total nitrogen limits), and goals for
sustainability, operability, and flexibility. Designers should work closely with facility owners and other
stakeholders to develop design criteria that carefully consider future needs and treatment conditions for
the entire planning period (typically 20 years). The text box on the next page highlights some key
questions that should be asked during design criteria development.
Design criteria are an important part of an overall facility plan (also called a strategic or master
plan). The facility plan documents the planning and decision-making processes that lead directly to
design and construction. Among other things, it identifies the problem, presents criteria and
assumptions, provides preliminary layout and cost alternatives, and reviews staff requirements. Ten
State Standards (2004) requires that comprehensive facility plans be completed for projects involving
new, upgraded, or rehabilitated wastewater treatment facilities. Facility plans are usually prepared by a
consultant and signed and sealed by a registered professional engineer. Additional guidance on
developing an overall facility plan is provided in the Water Environment Federation (WEF) Manual of
Practice No. 28, Upgrading and Retrofitting Water and Wastewater Treatment Plants (2005).
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Key Questions to Consider when Establishing Design Criteria
• Will the service area boundary change in the future? A planned residential community or
annexation of an existing community can significantly change influent flows and characteristics.
• What is the population projected to be at the end of the design period (typically 20 years)?
Population projections should consider seasonal changes in population and non-permanent
residents.
• Is decentralized or onsite treatment being encouraged for new, remote communities? Some
states and large wastewater utilities are encouraging decentralized treatment to reduce the
load on the existing collection system and save energy.
• Are there any anticipated changes in industrial users? Addition of a major industry or industrial
plant shutdowns can have a significant impact on wastewater quantity and quality.
• Is there a strong water conservation program in your service area or is one being planned?
Water conservation could result in reduced flow volumes but higher concentrations of nitrogen
and phosphorus in the wastewater.
• How certain are you about future regulations and flow quantity? Flexibility in design can help
address uncertainty in future conditions.
This chapter provides technical guidance on establishing design criteria for the purposes of
upgrading an existing facility to provide nutrient removal (nitrogen, phosphorus, or both) or improve
nutrient removal capabilities if they already exist. It is important to note that the data gathering steps in
Sections 7.2 through 7.4 go hand in hand with the development and calibration of a process simulation
model. As will be explained in Chapter 10, modeling is the recommended approach for designing
WWTP upgrades for biological nutrient removal (BNR) because of (1) its flexibility in enabling designers
to quickly test many different configurations and operating scenarios and (2) its ability to simulate
treatment performance under a wide range of conditions using dynamic modeling.
Many states have specific requirements for design criteria such as design flows and reliability.
Designers should carefully review state requirements and guidelines when developing criteria for
nutrient removal upgrades or retrofits. In addition to this chapter, readers are encouraged to review
Chapter 5, Upgrading Existing Facilities, in the Municipal Nutrient Removal Technologies Reference
Document (USEPA 2008), available online at http://www.epa.gov/OWM/mtb/mnrt-volumel.pdf.
General information on WWTP design can also be found in Ten State Standards (2004), WEF and ASCE
(1998), and Tchobanoglous et al. (2003).
7.2 Characterizing Existing Treatment
Technology selection and design for efficient nutrient removal requires an extensive
understanding of the existing plant's configuration, operation, and performance. For example, an
operations team wishing to optimize the plant's activated sludge system for nitrification needs in depth
knowledge of flow configuration, hydraulic residence time (HRT), solids residence time (SRT), and
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dissolved oxygen levels within the basin(s) during both normal and peak flow conditions. Recycle
streams, such as supernatant from sludge dewatering operations, can increase ammonia in the activated
sludge basins and have a negative impact on nitrification. The configuration of existing sludge handling
systems can be a driving factor in the design of chemical phosphorus removal systems. Odor control
practices in the collection system can reduce the amount of substrate available for biological
phosphorus removal.
In addition to configuration and performance, it is important to ascertain the condition of
existing equipment such as piping and pumps to determine their remaining life and rehabilitation
requirements. WEF (2005) recommends plant walk-throughs and interviews with the owners and
operators to gather this information and determine how existing equipment would fit into the upgraded
facility plan.
7.3 Design Flow Rates
The total flow delivered to a WWTP can vary from year to year due to population and industrial
changes within the service area. Influent flow rate can also vary seasonally—especially for areas with
large seasonal populations such as resort communities. WWTPs can experience significant peak flows
during wet weather events due to inflow and infiltration or if they have a combined collection system.
For small plants, peak hourly flow based on individual usage patterns can have a significant impact on
design.
Retrofits for nutrient removal should be designed for average and peak flow conditions. These
flow conditions should be projected for initial operation and through the end of the design period. The
typical design period is 20 years; however, shorter alternative design periods (e.g., 10 years) may be
used for the phased implementation of retrofits. Energy efficiency should be considered for existing
operation and at the end of the design period.
This section discusses the various design flow rates that are important for designing upgrades or
retrofits for nutrient removal.
7.3.1 Characterizing Existing Flow
Designers should use historical operating data from the plant to identify the various
components of the existing flow and to characterize its variability over time. At least three years of
historical flow data should be analyzed if available. Graphical and statistical analysis of this data can be
extremely useful. Flow characterization is also an important component of developing and calibrating a
process simulation model. Designers may decide to collect detailed flow data for modeling purposes.
Table 7-1 provides a description of influent flow components. Flow from domestic sources is a
key component that provides the basis for flow projections. Infiltration and inflow (l&l) and storm water
flow are needed for defining peak flow values. Table 7-2 lists the flow characteristics that should be
identified based on existing plant records.
Existing flow data should be used to identify a peaking factor for the plant. The peaking factor
is the peak flow rate (typically peak hourly flow) divided by the average long-term flow rate. At least
three years of flow data should be used to determine the peaking factor (Tchobanoglous et al. 2003).
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Table 7-1. Influent Flow Components
Flow component
Description
Estimation Methodology
Domestic wastewater
Wastewater from residential, commercial,
institutional, and similar facilities.
Average dry-weather flow minus
industrial flow. Can be estimated
assuming a typical flow per day per
person.
Industrial wastewater
Wastewater originating from industrial
processes. Characteristics are highly
variable. Can contain components that
inhibit biological nitrogen removal.
From water utility meter records or
interviews with owners.
Infiltration and inflow
(l&l)
Stormwater, snowmelt or ground water
that enters the collection system either
through cracks and other types of breaks in
the system (infiltration) or direct storm
drain connections such as foundation and
basement drains (inflow). Much lower
contaminant concentrations compared to
domestic wastewater.
l&l can be estimated as the difference
between average daily wet weather
flow and average daily dry weather
flow. Infiltration can be determined
separately from inflow by comparing
flow during wet periods when the
ground water is elevated, to flow
during and shortly after a heavy
rainfall.
Storm water
A factor only for localities that use the
same system for collecting wastewater and
storm water(combined systems).Infiltration
from peak rainfall is included in l&l above.
Difficult to differentiate between
inflow and storm water flow. Can be
estimated as the difference between
dry weather flow and wet weather
flow.
Table 7-2. Flow Characterization
Flow Characterization
Annual average daily
flow1
Average daily dry-
weather flow
Average daily wet-
weather flow
Maximum or peak
daily flow
Maximum monthly
flow
Peak hourly flow
Peak instantaneous
flow
Diurnal flow pattern
Atypical events
Description
Average flow rate occurring within a 1-day period based on annual flow data, i.e. l/365th
of total Annual Flow.
Average of daily flows sustained during a dry weather period. Determined as the daily
average flow for all dry weather periods during an average year.
Average of daily flows sustained during a wet weather period when l&l and storm water
are factors. Determined as the daily average flow for all wet weather periods during an
average year.
The largest volume of flow occurring within a 1-day period during an average year.
The average of daily flows for the month with the highest total flow during an average
year. Once determined should be used for design flow.
The largest volume of flow occurring within a 1-hour period during an average year.
Highest recorded flow rate. May be considerably below the actual peak flow because of
equipment limitations, but should be estimated if so.
The typical daily flow pattern for domestic wastewater, i.e. with peaks occurring in the
morning and evening.
Types of events include storm flows and large industrial discharges. Based on historical
records going back as many years as possible (e.g., based on review of past 10 years).
1. An average year is usually the average of three years, or more, at the discretion of the designer. However, its definition
may change depending upon whether or not there is a steady annual increase or decline for known reasons.
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7.3.2 Projecting Future Conditions
The following information is needed to project future flows and establish design flows for the
project:
• Projected population growth for the design period including the estimated number of people to
be served at the end of the design period.
• Per person (per capita) flow rate estimate in gallons per day (gpd).
• Projected commercial growth (including zoned areas and expected types of businesses) and new
institutional and recreational facilities.
• Planned industrial contributions and an allowance for unplanned contributions.
Population growth projections are often made by the city or county government. If these are
not available, historical population data can be extrapolated to predict population increases into the
future, although this method is crude at best.
Historical flow records can be used to estimate average per person flow for a given community
by dividing the average annual dry-weather flow (expressed as a daily value) by the number of people
served. Industrial flows should be subtracted from the average annual dry-weather flow if possible.
Historical records, however, cannot always be depended upon to give an accurate prediction of future
flows. In its onsite wastewater treatment manual, EPA (2002) reports that average daily flows between
50 and 70 gpd per person are typical for residences built before 1994. New homes and residences
constructed after 1994 are subject to the U.S. Energy Policy Act (EPACT) standards, which set maximum
flow values for household plumbing fixtures such as toilets, showerheads, and faucets (Table 7-3 shows
the difference in water use between standard fixtures and EPACT fixtures). EPA estimates that homes
built after 1994 or homes that have been retrofitted with EPACT standard fixtures have average daily
wastewater flows between 40 and 60 gpd per person (EPA, 2002). Average per capita flow may be even
less in regions that have enacted water conservation programs.
Table 7-3. Comparison of Flow Rates and Flush Volumes Before and After U.S. Energy Policy
Act
Fixture
Kitchen faucet
Lavatory faucets
Showerheads
Toilet (tank type)
Toilet (valve type)
Urinal
Fixtures installed
prior to 1994
3.0 gpm (0.19 L/s)
3.0 gpm (0.19 L/s)
3.5 gpm (0.22 L/s)
3.5 gal (13.2 L)
3.5 gal (13.2 L)
3.0 gal (11.4 L)
EPACT requirements
(effective January,
1994)
2. 5 gpm (0.16 L/s)
2. 5 gpm (0.16 L/s)
2. 5 gpm (0.16 L/s)
1.6 gal (6.1 L)
1.6 gal (6.1 L)
1.0 gal (3.8 L)
Potential reduction
in water
used (%)
16
16
28
54
54
50
Source: Konen, 1995.
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Design textbooks also provide typical ranges of flow per person per day. WEF and ASCE (1998)
provide a range of total flow per capita of between 60 and 110 gpd/person. Tchobanoglous et al. (2003)
present a possible range of per person flow rate values of between 46 and 97 gpd based on the number
of persons per household, with higher per person values for smaller households (citing Mayer 1999).
Some states provide guidance on selecting per capita flow rate estimates. The Wisconsin Department of
Natural Resources (DNR) recommends that design engineers assume 60 to 70 gpd/person for plants
serving 5,000 or fewer people, and between 65 and 80 gpd/person for plants serving more than 5,000
people. This higher value reflects greater influence of commercial and industrial customers (Wisconsin
DNR 2006). When completing flow projections, regional and state guidance should be consulted to
account for the potential regional differences in water use and wastewater generation.
Water use records are another source of data that can be useful in predicting per capita flow
and future flows. The 1999 report, Residential End Uses of Water (Mayer 1999), is one of the most
commonly cited sources for water use data. The proportion of water consumed that becomes
wastewater depends on regional conditions. Higher proportions (as high as 90 percent) can be assumed
for northern states during cold weather when irrigation use is low. A smaller proportion (60 to 70
percent) becomes wastewater in arid regions (Tchobanoglous et al. 2003).
Future flow projections for new commercial, institutional, and recreational facilities can be
based on meter records from similar facilities. Data Tables 3-2 through 3-4 in Tchobanoglous et al.
(2003) provide typical values if flow data are not available.
7.3.3 Setting Design Flow Rates
Average annual design flow is important for conducting steady state modeling simulations. It
should be based on population projections and estimated future flows from commercial, institutionAL,
recreational, and industrial facilities for the end of the design period (or at some point in the design
period if using a phased approach as described in Section 7.3). Average annual flow should represent
normal, dry-weather operating conditions.
Biological reactors (e.g., activated sludge basins) are usually sized using the Maximum Month
Design Flow, although short SRT systems may require sizing based on max week or even short running
averages. Clarifiers, filters, and pump stations are typically sized based on maximum hydraulic or solids
load, which is often based on Peak Hourly Design Flow. Maximum and peak rates are usually derived
by multiplying the average annual design flow by a peaking factor. Peaking factors are based on analysis
of existing peak flows compared to average flows (see Section 7.3.1), with engineering judgment applied
to account for future changes in l&l (e.g., collection system upgrades can significantly reduce l&l, but
aging infrastructure increases l&l). Flow equalization is a significant factor—flow equalization at the
plant or in the collection system can reduce the peaking factor. Diurnal flow pattern for the design year
are also very important for modeling.
Designers should consult with state regulators to ensure that design flows adhere to state
requirements and meets the forecasted needs of the community through the end of the design period.
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7.4 Wastewater characteristics
Measured data characterizing the quality of wastewater in the influent and at key locations
throughout the plant are extremely important for designing biological nutrient removal processes.
Mathematical models depend on accurate data to predict the performance of various treatment
scenarios. The phrase "garbage in, garbage out (GIGO)" is very appropriate here - model predictions are
only as good as the data on which they are based. As will be explained further in Chapter 10, one of the
key benefits of modeling is the ability to simulate variations in wastewater flow and characteristics and
thus, produce a "tighter" design with lower process safety factors, saving on capital and operations
costs. Tight designs with low safety factors are highly dependent on good quality data. To address the
importance of wastewater characterization, the basic guidance on data collection in Section 7.4.1 is
followed by recommended steps for data verification in Section 7.4.2.
It is very important to note that characterization of influent flow rates and mass loading of
contaminants go hand in hand. A common error in design is to assume that nutrient loadings are
independent of flow rate variability, which is not the case in systems that receive any l&l or storm water
flow. Conversely, nutrient loadings from domestic sources often vary diurnally and can be greatest
during periods of high flows. As noted throughout this document, return streams can discharge high
nutrient concentrations to the head of the treatment plant
7.4.1 Data Collection
Which data are typically available from the plant?
Plants routinely measure flow and wastewater quality to aid in basic operating decisions and to
ensure regulatory compliance. Most maintain historical data in electronic format such as spreadsheets,
databases, or, in the case of small systems, handwritten log sheets. Daily composites or grab samples of
the influent and effluent are typically available from plant personnel for the following parameters
(Melcer et al. 2003): Biochemical Oxygen Demand (BOD)1, Total Suspended Solids (TSS), mixed liquor
suspended solids (MLSS), ammonia and/or Total Kjeldahl Nitrogen (TKN), nitrate, and phosphorus
(sometimes only ortho-phosphorus).
Designers should start by obtaining as much historical data as possible from the wastewater
utility. When working with large plants with electronic records, designers should obtain as many as 8 to
10 years of historical data to help identify annual trends and eliminate data anomalies and errors. It is
important, however, to only include data from time periods that represent current conditions. Dates of
major plant configuration or operating changes should be discussed with plant operators and
considered when reviewing historical data. Sudden changes in influent conditions could also represent
addition or removal of a major industry or a change in laboratories.
1 For the purposes of this design manual "BOD" represents the 5-day BOD measurement method (sometimes
referred to as BOD5) unless otherwise noted.
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What additional data are needed for modeling nutrient removal processes?
In general, the following data are needed to model nutrient removal processes:
1) Average values for the following parameters in the treatment plant influent, primary effluent,
and secondary effluent.
• Chemical oxygen demand (COD) - COD is used instead of BOD to model organic substrate;
see the text box in this section for further explanation. If the
• Total suspended solids (TSS)
• Volatile suspended solids (VSS)
• Total Kjeldahl Nitrogen (TKN, which is the sum of organic and ammonia nitrogen)
• Ammonia-nitrogen (NH4-N, NH3-N )
• Nitrate-nitrogen (NO3-N)
• Total phosphorus (TP)
• Orthophosphate (OP)
• Alkalinity
• Temperature
• pH-spot samples. Average of samples is meaningless
2) Diurnal influent loading patterns for COD, ammonia, and phosphate are also very important for
both steady state and dynamic modeling.
Data should represent normal dry-weather operating conditions and be based on several years of
historical data.
In most cases, historical data are insufficient to meet design objectives for nutrient removal. If
time and funding allow, measurement campaigns, also called "special sampling," can be conducted early
in the process before calibration begins. This step is recommended if key data are missing or if
designers do not have confidence in existing data (e.g., grab sampling locations are not representative,
historical data do not match expected values). The duration of a typical measurement campaign is 1 to
2 weeks. 24-hour composite samples should be collected from well-mixed locations representing
influent flow, primary effluent, and secondary effluent and analyzed for key parameters such as COD2,
BOD, VSS, TSS, ammonia, nitrate, phosphate, alkalinity, and pH. Designers should consider
supplemental sampling of the influent to obtain diurnal profiles (e.g., hourly grab samples for 24 to 48
hours measuring COD, ammonia, and phosphate). Special sampling of operating parameters could
include spot checks of dissolved oxygen (DO) concentrations in aerated zones, daily grab samples of
mixed liquor, monitoring of RAS recycle rate and concentration, and sampling to characterize side
streams (Melcer et al. 2003). Where ample historical BOD information and no COD data are available,
designers should try to establish a relationship between the COD and BOD for raw and settled
wastewater. See the text box at the end of this section for guidance on estimating COD if only BOD data
are available.
Total cost of a 2-week composite sampling campaign is estimated to be about $20,000 (Takacs 2009)
2 Laboratory analysis for COD should be done as soon as possible after sample collection to minimize possible
errors introduced by preservation techniques (Takacs 2009).
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Why do mathematical models use COD instead of BOD to represent organic material?
To explain why COD is the recommended measurement for organic material in wastewater, it is
useful to first provide a review of the way in which microorganisms consume organics.
Microorganisms use organic substrate in two ways.
(1) They oxidize it to CO2 and water to provide energy for maintaining existing cell mass.
This is accomplished by transferring electrons from the organic substrate to a
terminal electron acceptor (oxygen, or in the case of anoxic conditions, nitrate).
Under substrate limiting conditions, which are typical for activated sludge systems,
microorganisms use a relatively fixed percentage of organic substrate, expressed as
COD, for the cell maintenance process.
(2) They convert organic material to new cell mass using the energy generated through
oxidation, as described in item (1) above.
COD measures the amount of an oxidant that reacts with a sample under controlled
conditions. Because each mole of oxygen accepts four electron equivalents, the COD
measurement of oxidant demand is a direct measure of the electron donating potential of a
sample. From the electron donating potential, it is possible to quantify the amount of COD and
oxygen consumed and the amount of sludge produced.
The BOD measurement method has several limitations. It only accounts for the
carbonaceous material that is used for energy, not organic material converted to new cell mass.
Most BOD measurements are based on a testing duration of 5 days (BOD5) which only accounts
for the amount that reacts within a five-day period and not all the available carbon. Depending
on how rapid the oxidation rates are, two different wastewater samples may have the same
ultimate BOD but different BOD5 concentrations.
Total organic carbon (TOC) could be used to represent carbon; however, because models
account for reactions based on the electrons transferred and not the moles of carbon, it would
be difficult to track the reactions based on TOC.
Measuring COD at various locations in the wastewater plant allows for a mass balance.
COD entering the system must equal the sum of (1) COD in the effluent, (2) COD of the wasted
sludge, and (3) oxygen consumed in utilization of organic material.
For more information: See Melcer et al. (2003), Appendix B; APHA, AWWA, and WEF (2005).
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What if the plant does not have COD data? This is a common problem. Many plants measure BOD
instead of COD because the BOD method has been used for a long time and is often the basis for
regulatory compliance. If COD measurements are not available, COD can be approximated by
establishing the ratio of COD/BOD or COD/VSS. While some mathematical models may take BOD as
the input, it is still necessary to adjust the COD fractions for good correlation.
Although measured values of BOD, COD, and VSS in the plant influent vary greatly, the ratios are
generally consistent for a specific wastewater. COD/BOD ratios are typically 1.9 to 2.2, with the higher
ratios for relatively fresh wastewater. COD/ VSS should range from 1.42 to 1.48 mg COD per mg
MLVSS. Estimating the ratio of COD/BOD is more common than COD/VSS.
To determine the COD:BOD ratio for a specific wastewater:
• Collect several samples (10 to 20 are recommended) of influent wastewater. The samples
should be representative of total influent flow and taken at different times of the day at low
and high influent flow rate.
• Split the samples, taking care to homogenize them thoroughly.
• Analyze each sample for COD and BOD. Use the same method for BOD that is used for the
available plant data. (BOD5 is most common, although some plants will have cBOD5 data.)
• Determine the average COD:BOD ratio for all samples.
• Repeat the test with primary effluent
Designers can then apply this ratio to historical BOD data to determine COD.
Limitations in funds and time, however, often render extensive measurement campaigns
impractical. Melcer et al. (2003) recommends an iterative, or "tiered" approach whereby the model is
first run using historical data. Additional data needs are determined based on the accuracy of mass
balances and other calibration checks. See Chapter 10 for detailed guidance on model calibration using
this recommended iterative approach.
7.4.2 Data Verification
The designer should perform a number of data review and verification steps before making
calculations or entering data into the model. In the simplest terms, designers should critically evaluate
the data to answer the questions, "do the data make sense?" and "do they follow expected engineering
principles?" Below is a recommended 4-step process for reviewing and verifying data. Chapter 10
provides additional recommended data verification steps for organic substrate, nutrient fractions, and
kinetic/stoichiometric parameters, respectively. Additional mass balances and calibration checks are
recommended in Chapter 10 of this design manual.
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Step 1: Review data for anomalies.
• Conduct statistical analyses to determine average, median, minimum, and maximum values.
Look for obvious outliers that could be a result of data entry error (e.g., a parameter different
than comparable measurements by a factor or 10 could be an error in decimal placement).
• Plot historical data over time to evaluate trends. Extreme changes in readings or flat data could
indicate sensor problems.
Step 2: Review sampling locations and collection procedures.
• Permanent instrumentation should represent the process and should be in a well-mixed
sampling location.
• Check locations of flow meters—are separate flow meters installed for each train? If not, how is
split estimated?
• Check location of grab sample collection—sometimes a convenient place to sample from a basin
is not representative of the water quality in that basin.
Example 1: Improper Sampling of Primary Effluent.
An automated flow-proportional composite sampling device was drawing sample aliquots from near
the bottom of a primary effluent channel. The primary effluent channel was not well mixed and thus,
the sampling device was drawing an unrepresentative amount of particulate matter. A second
problem existed with the sampling device itself. The vacuum used to withdraw and transport the
sample from the primary effluent channel to the sample container caused the volatile portion of the
organic matter to dissipate as it entered the receiving vessel (Melcer et al. 2003).
Step 3: Review analytical procedures.
• Check that samples are analyzed using approved standard methods.
• Check the method for determining BOD. Because modeling experience is based on relationships
between COD and BOD, all data should be from the 5-day BOD test (also referred to as BOD5),
not ultimate BOD. Check if BOD is in the form of carbonaceous BOD (cBOD). Plant managers
are more often monitoring and reporting cBOD instead of BOD to be consistent with state
permit limits. The standard method for cBOD requires the addition of an inhibitor to prevent
nitrification that would otherwise register as BOD. There is evidence that some of the organic
reactions are also inhibited, leading to artificially low cBOD values (Albertson 1995). Others
(Baird et al. 2002) maintain that there is no inhibition of BOD but, instead, there are more
nitrifiers in raw wastewater than usually assumed, leading to an apparent inhibition of BOD. To
check, determine the COD/BOD ratio if COD data are available. The ratio should be between 1.9
and 2.2. COD/BOD ratios higher than 2.2 may indicate inhibition for cBOD data or could result
from excessive particulate matter from industries. Some designers have corrected this by
adjusting cBOD data upward by 10 to 15 percent for domestic wastewater.
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• Check the method for COD determination. Mercuric sulfate (HgSO4) must be added to counter
the influence of chlorides, and silver sulphate A(gSO4) is required as a catalyst to facilitate for
the oxidation of some organic matter such as VFAs. These chemicals are hazardous and must be
disposed of following regulations. The use of mercury salts is necessary when chloride is a
factor in COD, but can be omitted in instances where the contribution of chloride to COD is
negligible. When using historical COD values, check to see whether these chemicals were used.
COD/BOD ratios should be as expected (see Step 4). See Standard Methods for the Examination
of Water and Wastewater (APHA, AWWA, and WEF 2005) for more information about methods
5220Athrough D.
Example 2: Significant Observed Differences in COD in Split Samples
Latimer et al. (2007) reported several cases of questionable wastewater characterization data. In
one instance, large differences in COD concentration were observed between split samples analyzed
on-site and at an approved commercial laboratory. Although both used EPA approved methods and
typical QA/QC procedures, the laboratory results were on average approximately 30 percent less
than plant values. This discrepancy persisted even after efforts to ensure proper sampling,
preservation, and homogenizing of the sample prior to splitting. The only explanation provided by
the laboratory was that the method gave different results depending on the way the vial was turned
in the instrument. This example underscores the importance of following the manufacturer's
instructions with respect to homogenization of samples and to the correct orientation of vials in the
spectrophotometer. There are many cases similar to this example, and if the COD information looks
suspect when modeling, consider that the method used to determine it may be at fault.
• Check the size of the filter used to differentiate between TSS and DSS (typically 1.2 to 1.5 u.m
glass fiber filters should be used). This becomes important when evaluating fractions.
• Check that operators used appropriate sample preservation and storage techniques and met
maximum allowable holding times.
• Check the quality assurance procedures. Are duplicates and blanks regularly sent to the
laboratory?
• Check that sensors and other measurement equipment are regularly calibrated. For example,
Most TSS meters should be calibrated at least weekly.
If membrane electrodes are used for DO measurement, they should be replaced at least
quarterly.
See Chapter 13 on instrumentation and controls for additional recommendations.
Nutrient Control Design Manual 7-12 August 2010
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Example 3: Error in Calculation of Totalized Flow
At one treatment plant, comparison of instantaneous flow measurements to totalized flow showed
large differences. An investigation revealed that the programmable logic controller (PLC) calculating
the total flow contained an error, resulting in an underestimate of total plant inflow of approximately
25 percent (Third et al. 2006). In another instance, the flow meter was metering only one of two
parallel streams. The PLC would multiply this by a factor of two for total plant flow. However, for
the first several years, only one of the two flow trains was used and the flow was reported at double
the actual flow.
Step 4: Compare measured ratios to typical values for domestic wastewater.
• COD/BOD: The COD/BOD ratio for influent wastewater that has a low detention time in the
collection system (i.e., fresh wastewater) should be 2.2 or higher. Wastewater that travels
through fairly flat collection lines and/or force mains ferments, which increases the BOD5 but
not the ultimate BOD and thus reduces the COD/ BOD ratio to around 2.0 or less. The COD/
BOD ratio will be higher when the VSS/ BOD ratio is higher.
• COD/VSS ratio: This ratio should be 1.42 to 1.48 mg COD per mg MLVSS.
• VSS/TSS ratio: This ratio should be on the order of 0.75 mg VSS/TSS for municipal wastewater
and 0.83 mg VSS/TSS for primary effluent.
• COD/TKN ratio: Although TKN and COD concentrations vary in the plant influent, the ratio
should be fairly constant. For raw domestic wastewater, the ratio varies from approximately 9.5
to 12.
• The COD/TP ratio: Values should range from 35 to 60.
• TKN/VSS ratio: This ratio should vary between 0.08 and 0.1 for mixed liquor VSS but may vary
much more for raw domestic wastewater.
• Sludge generation (measured as mass VSS or TSS generated) per mass COD or BOD in influent:
Check against similar plants with similar SRTs. Figure 7-1, which shows sludge production with
and without primary treatment for different temperatures and SRTs, can be used as a guide.
• The TSS and BOD concentrations should be within 10 percent.
Nutrient Control Design Manual 7-13 August 2010
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primary treatment (Ib/lb = kg/kg).
Source: WEF and ASCE 1998, Figure 11.7, p. 11-20. Reprinted with permission from Design of Municipal Wastewater
Treatment Plants - MOP 8, 4th Edition, Copyright ©1998, Water Environment Federation and American Society of Civil
Engineers, Alexandria, VA. www.wef.org
Nutrient Control Design Manual
7-14
/August 2010
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7.5 Target Effluent Concentrations for Total Nitrogen and Total Phosphorus
Plants should base their target effluent concentrations for total nitrogen (TN), and total
phosphorus (TP) on the following:
• Nutrient limits in National Pollutant Discharge Elimination System (NPDES) permits including
numeric limit and averaging period requirements
• An overall process design factor
In other words, the design value should be below the permitted value to provide a margin of safety.
Issues related to NPDES permit limits are discussed in this section. Application of an overall process
design factor (i.e., safety factor) is addressed in Section 7.9
NPDES permits are set on a 5-year cycle. Historically, NPDES permits sometimes included an
effluent limit for ammonia-nitrogen but not TN or TP. Permits that included total nutrient limits often
set maximum allowable values of 10 mg/L for TN and 1.0 mg/L for TP. Recognizing the harmful effects
of eutrophication of the nation's waterways due to excessive nutrient loadings (see Chapter 2, Section
2.4 of this manual for a complete discussion), regulatory agencies have started setting lower nutrient
limits in NPDES permits. TN limitations of 5.0 or 3.0 mg/L and TP limits of 0.5 and even 0.1 mg/L or less
are becoming much more common.
Particularly in western portions of the country, nutrient-related water quality problems can
occur on a seasonal basis. As a result, water quality-based effluent limitations for nutrients usually
apply only during the "critical period." This period routinely occurs from late spring through early fall
when stream flows are low and temperatures are warm.
The averaging period for the permit limits is an extremely important factor in selecting
technologies, establishing design parameters, and operating plants for nutrient control. The following
approaches to limiting effluent nitrogen and phosphorus have been used in NPDES permits, either alone
or in combination:
• Annual average
• Quarterly average
• Maximum month
• Maximum week
• Maximum day
• Seasonal average
Table 7-4 provides several examples of NPDES permit limits for nutrients as presented during a
2008 workshop on nutrient removal. Other examples can be found in state records and throughout the
related literature, and individual NPDES permits and fact sheets can be accessed through Envirofacts.
Visit the EPA webpage http://cfpub.epa.gov/npdes/permitissuance/permitscanning.cfm for more
information.
Nutrient Control Design Manual 7-15 August 2010
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Table 7-4. Example Permit Limits for Nutrients
WWTP
Name
Truckee Meadows
Water
Reclamation
Facility
River Oaks
Advanced
Wastewater
Treatment Plant
Fiesta Village
Advanced
Wastewater
Treatment Plant
Orange County
Eastern Water
Reclamation
Facility
Parkway
Wastewater
Treatment Plant
Clark County
Water
Reclamation
District
Rock Creek
Advanced
Wastewater
Treatment Facility
Blue Plains
Advanced
Wastewater
Treatment Facility
Location
Reno, NV
Tampa, FL
Fort Myers,
FL
Orlando, FL
Laurel, MD
Las Vegas,
NV
Washington
County, OR
Washington,
DC
Ann.
Avg.
Flow
(MGD)
10
5
17.3
5.5-6
110
30 (dry
weather)
370 (dry
weather)
TN Limits
mg/L
2.0 daily max for
nitrate
6.00 daily max
4.50 wkly avg
3.75 monthly avg
3. 00 annual avg
6.0 single sample
4.5 wkly avg
3.0 monthly avg
3.0 annual avg
6.0 wkly avg
5.0 monthly avg
3.0 annual avg
7.0 monthly avg
and 11.0 wkly
avg between 4/1
and 10/15
0.6 mg/L as
ammonia
None provided
7.5 annual avg
changing to 4. 2
annual avg in
future
Lbs/
day
500
Silas
ammo
nia
TP Limits
mg/L
0.4 30-day avg
2.00 daily max
1.50 wkly avg
1.25 monthly avg
1.00 annual avg
1.0 single sample
0.75 wkly avg
0.50 monthly avg
0.50 annual avg
2.4 weekly avg
2.0 monthly avg
1.0 annual avg
1.5 weekly avg
1.0 monthly avg
0.2 mg/L
0.10 mg/L
monthly median
0.35 monthly avg
0.18 weekly avg
Lbs/
day
134
176
Citation
Gray 2008
Phillips 2008
Meyer 2008
Madhanagopal
2008
Selock 2008
Drury2008
Spani 2008
Baily2008
Source: WEFTEC Technical Education Workshop No. 101: WEF/WERF: Demonstrated Processes for Limit of Technology
Nutrient Removal: Achievable Limits and Statistical Reliability. Chicago, IL. October, 2008.
7.6 Goals for Reliability, Sustainability, and Process Flexibility
Reliability is the capability of a treatment process to perform its intended function without
failure or interruption of treatment such as during floods, power failures, or in the event of equipment
failure. Reliability goals of a nutrient removal process should depend on the permit limits and averaging
period. The most common safety factor is to design for maximum month loadings. Reliability is more
important and may necessitate larger safety factors if the average period is daily or weekly. Monthly,
seasonal or yearly permit levels provide some degree of flexibility for unforeseen events without
exceeding the permit level. See Chapter 6, Section 6.7 for additional discussion on reliability of
technologies and effluent limits.
Nutrient Control Design Manual
7-16
August 2010
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It is common for states to require that treatment components be designed to treat the average
daily design flow with the largest unit out of service. Ten State Standards (2004) requires that all plants
be provided with an alternative source of electric power or pumping capacity to allow continuity of
operations during power failures. Methods of providing alternative power include connecting to at least
two independent power sources and using backup power generators. Spare parts should be kept on site
to ensure continuous nutrient removal capabilities, and equipment should be accessible for repairs.
Designers should strongly consider incorporating online instrumentation and controls into
design. Online instrumentation can provide rapid feedback compared to grab sampling and enable
operators to respond more quickly to environmental changes, thereby maintaining nutrient removal to
desired levels. Automated control can significantly reduce workload while at the same time improving
control of process parameters—thus improving treatment plant performance, often at reduced energy
costs. Automated control is highly recommended for plants trying to achieve low effluent nutrient
levels (e.g., TN < 3 mg/L or TP < 0.1 mg/L). See Chapter 13 of this manual for guidance on online
instrumentation and controls.
Sustainability implies a simultaneous focus on economic, social, and environmental performance
(USEPA 2008). One growing area of focus in recent years is the inter-relationship among water, energy,
and greenhouse gas emissions. In its recently published document Ensuring a Sustainable Future: An
Energy Management Guidebook for Wastewater and Water Utilities, EPA notes that energy
management is at the heart of efforts to ensure sustainability for utilities (USEPA 2008b). When
considering upgrades and retrofits to existing WWTPs for the purposes of removing nutrients, designers
should work with owners to identify sustainability goals such as minimizing energy use, resource
recovery, and reducing greenhouse gas emissions.
Flexibility in design can help address uncertainties in future conditions. For example, the design
should include the flexibility to increase its treatment capacity if owners and regulators suspect that
nutrient limits will decrease in the future based on new watershed analyses and the setting of numeric
water quality criteria. Similarly, the design should address the elements of uncertainty in population
projections and assumptions regarding the industrial and commercial components of wastewater. The
changes in influent flow as a result of water conservation have been studied (AWWARF 1999) but are
very area-specific. One method to address uncertainty in the design includes dividing projects into
several phases to ease the financial burden on wastewater utilities. This phased approach also reduces
uncertainty by meeting shorter-term design objectives first rather than constructing facilities to meet
projected criteria through the end of a 20-year planning period. Turndown capacity of blowers has
been found to be critical in the design of energy-efficient aeration systems and often compensates for
protective design assumptions. WEF (2009) recommends a minimum turndown ratio of 5:1. Multiple
blowers of more than one size also may be installed so that blowers can be turned off when possible.
Automation of turnoffs has proven to improve energy efficiency.
Operational flexibility is often needed to remove nutrients to low levels in areas with significant
cold weather periods. Such flexibility may also be advantageous for plants with highly variable influent
flow and no flow equalization. One way to build operational flexibility into design is to include swing
zones. Swing zones can be located between the anaerobic and anoxic zones, the anaerobic and aerobic
zones, or the anoxic and aerobic zones depending upon the secondary processes selected (USEPA 2008).
For example, a swing zone between the anoxic and aerobic zone would contain both mixers and
diffusers. If more SRT were required for nitrification (e.g., during winter months), the diffusers could be
Nutrient Control Design Manual 7-17 August 2010
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operated. During warmer months, the diffusers could be turned off to create anaerobic conditions and
promote phosphorus removal.
Example 1: Successful Use of Swing Zones to Remove Phosphorus under Changing Conditions.
Vackson et al. (2007) report successful use of swing zones to provide phosphorus removal at a 5 MGD
plant at The Colony, Texas. Based on new total nitrogen permit limits, the plant operation team had
decided to retrofit their existing activated sludge process with an Integrated Fixed-Film Activated
Sludge (IFAS) system and modify their existing contact stabilization process for plug flow. After the
design was finalized and construction was about to begin, the team learned of a pending revision of
their phosphorus effluent limit to 1 mg/L The design was modified to include two anaerobic swing
zones containing both coarse air diffusers and mixer systems. The authors report success in
achieving both nitrogen and phosphorus removal by operating the zone anaerobically in all but the
coldest months.
Safety and security are important considerations for any upgrade project. All new structures
should be enclosed and should have signs discouraging entry of anyone but authorized personnel.
Handrails and guardrails should be installed around tanks, trenches, pits, and other hazardous
structures. Warning signs should be appropriately placed, and adequate ventilation should be provided.
Designers should work with plant owners and other stakeholders to determine if additional security
features such as lights, motion detectors, and cameras are needed.
7.7 Sludge Treatment Options
Both the quantity and quality of the sludge produced should be considered when establishing
design criteria for nutrient removal processes. Biological nitrogen and phosphorus removal processes
do not generally produce additional sludge quantities. In fact, in the case of nitrogen removal operated
without the addition of an external carbon source, the biological nutrient removal processes will lower
sludge quantities. The addition of chemical treatment for phosphorus removal, however, always
produces additional sludge due to the generation of metal or calcium and phosphorus precipitates.
The addition of nutrient removal changes the sludge composition, which in turn can change the
thickening and dewatering properties. Nitrogen and/or phosphorus removal can improve thickening
characteristics due to decreased amounts of filamentous bacteria in the activated sludge. Sludge from
biological phosphorus removal processes will contain additional phosphorus, but will otherwise be
similar to sludge produced by conventional activated sludge systems. Sludge settling and dewatering
characteristics are generally the same or improved. Potential issues include phosphorus release and
struvite precipitation under anaerobic processing conditions. See Chapters 4 and 5 of this manual for
additional discussions of the effects of nitrogen and phosphorus removal, respectively, on sludge
handling.
7.8 Site Constraints
Site constraints can be a driving factor in selecting a nutrient removal technology. Particularly
for larger WWTPs in urban areas, space can be limited. Attached growth and hybrid systems such as
IFAS and moving bed bioreactors (MBBRs) can achieve nutrient removal in smaller footprints and can
Nutrient Control Design Manual 7-18 August 2010
-------
often be used to retrofit an existing basin without increasing the footprint at all. Chapter 6 and Chapter
8 provide additional discussion of footprint requirements for nutrient removal technologies.
Odor can be a serious issue for WWTPs, particularly those in urban areas. Designers should
always consider how a plant upgrade or retrofit for nutrient removal will impact the release of odors at
the plant. Containment (e.g., covering of treatment units) and treatment of captured gas should be
considered for each process upgrade. For example, the addition of fermentation to produce substrate
for biological phosphorus removal can produce significant odors. For this reason, sludge thickeners for
fermentation should be covered and the head space scrubbed to control odors. Many retrofits for
biological nutrient removal, however, do not increase odors. For example, the addition of anaerobic
zones for biological phosphorus removal does not increase odors because of the continuous addition of
oxidized activated sludge to the zone and also because volatile compounds such as H2S and volatile fatty
acids (VFAs) are not being stripped and the VFAs are used by the phosphate accumulating organisms.
Because improvements to WWTPs directly impact the community, the public should be brought
into the process as early as possible to educate them on the need for treatment changes and to obtain
their support. Designers and owners should consider the following options for public involvement (WEF
2005):
• Schedule a public meeting at the beginning of the planning process to identify major issues.
Schedule additional meetings throughout planning, design, and construction.
• Consider low-cost improvements such as tree planting and improved storm drainage to gain
public support.
• Establish a citizen's advisory group for large projects.
• Develop procedures for dealing with public complaints.
• Communicate with the public regularly through flyers, newsletters, and interviews on local news
stations. Regularly post project updates including pictures on the utility website.
7.9 Selecting an Overall Process Design Factor
Engineers routinely apply safety factors during design, particularly for sizing of basins and other
capital equipment, to account for uncertainties in predicted conditions and technology performance.
For design of nutrient removal systems, relevant uncertainties are:
• Variations in predicted influent flow rates and contaminant loadings.
• Uncertainty in predicted treatment performance (often driven by uncertainty in biokinetic
parameters).
• Variability in mechanical equipment.
As uncertainty in each of these factors increases, so does the safety factor. Note, however, that in
addition to inefficiencies in operation, oversized units also can result in reduced performance of
Nutrient Control Design Manual 7-19 August 2010
-------
biological nutrient removal systems, particularly those that include enhanced biological phosphorus
removal. For such system, the best safety factor is multiple trains rather than larger reactors. Then
trains can be taken on or off line to maintain near optimum performance. Additional time spent early in
the design process to carefully evaluate data and predict conditions throughout the design period can
allow for the use of a smaller factor and save money in construction costs.
Safety factors are highly dependent on the extent of data collection, calibration, and type of
process simulation modeling performed for the project. As confidence with model parameters increase,
the size of the safety factor decreases. One major advantage of dynamic modeling is its ability to
simulate a number of different future conditions and thus, enable the use of lower safety factors. See
Chapter 10 for additional discussion.
7.10 References
Bailey, W. 2008. Manager's Perspective on Multiple Step Chemical Addition (Blue Plains) for Low P.
Presented at WEFTEC Technical Education Workshop No. 101, Demonstrated Processes for Limit of
Technology Nutrient Removal: Achievable Limits and Statistical Reliability. WEF/WERF Chicago, IL.
October, 2008.
Bratby, J., J. Jimenez, D. Parker. 2008. Dissolved Organic Nitrogen - Is It Significant, and Can it be
Removed? Presented at WEFTEC 2008.
Drury, D. 2008. Manager's Perspective on Following BioP with Tertiary Chemical Removal Steps (Clark
County). Presented at WEFTEC Technical Education Workshop No. 101, Demonstrated Processes for
Limit of Technology Nutrient Removal: Achievable Limits and Statistical Reliability. WEF/WERF.
Chicago, IL. October, 2008.
Gray, R. 2008. Manager's Perspective on Separate Stage Fixed Film N and DN Processes for LOT
(Truckee Meadows Water Reclamation Facility). Presented at WEFTEC Technical Education Workshop
No. 101, Demonstrated Processes for Limit of Technology Nutrient Removal: Achievable Limits and
Statistical Reliability. WEF/WERF. Chicago, IL. October, 2008.
Jackson, D.R., LE. Ripley, T. Maurina, and S. Hubbell. 2007. Up to the Challenge : IFAS Helps Growing
Texas City Meet Discharge Limits, Expand Capacity. Water Environment & Technology. 19(11): 50-54.
Jeyanayagam, S. and I. Venner. 2007. Wastewater Process Design with Energy Savings in Mind. Florida
Water Resources Journal. January, PP. 28-33.
Konen, T. P. 1995. Water Use and Efficiency Under the U.S. Energy Policy Act. Stevens Institute of
Technology, Building Technology Research Laboratory. Hoboken, NJ.
Lancaster, C.D. and J.E. Madden. Not So Fast! The Impact of Recalcitrant Phosphorus on the Ability to
Meet Low Phosphorus Limits. Presented at WEFTEC 2008.
Leenheer, J., A. Dotson, and P. Westerhoff. 2006. Dissolved Organic Nitrogen Fractionation. Annals of
Environmental Science, Vol 1, 45-56.
Nutrient Control Design Manual 7-20 August 2010
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Madhanagopal, T. 2008. Manager's Perspective of a Combined Process (5-Stage Bardenpho) Process
for LOW N (Eastern Water Reclamation Facility) - Moderate Climate. Presented at WEFTEC Technical
Education Workshop No. 101, Demonstrated Processes for Limit of Technology Nutrient Removal:
Achievable Limits and Statistical Reliability. WEF/WERF. Chicago, IL October, 2008.
Mayer, P.W., W.B, DeOreo, E.M. Opitz, J.C. Kiefer, W.Y. Davis, and B. Dziegielewski. 1999. Residential
End Uses of Water. AWWAand AWWA Research Foundation.
Melcer, H., P.L Dold, R.M. Jones, CM. Bye, I. Takacs, H.D. Stensel, A.W. Wilson, P. Sun, and S. Bury.
2003. Methods for Wastewater Characterization in Activated Sludge Modeling. WERF Final Report.
Project 99-WWF-3.
Meyer, J. 2008. Manager's Perspective on Following Separate Stage Nitrification with Fixed Film
Systems for Low N (Fiesta Village). Presented at WEFTEC Technical Education Workshop No. 101,
Demonstrated Processes for Limit of Technology Nutrient Removal: Achievable Limits and Statistical
Reliability. WEF/WERF. Chicago, IL. October, 2008.
Mulholland, M.R., N.G. Love, D.A. Bronk, V. Pattarkine, A. Pramanik, and H.D. Stensel. 2009.
Establishing a Research Agenda for Assessing the Bioavailability of Wastewater Treatment Plant-
Derived Effluent Organic Nitrogen in Treatment Systems and Receiving Waters. February 27.
Mulholland, M.R., N.G. Love, V.M. Pattarkine, D.A. Bronk, and E. Canuel. 2007. Bioavailability of
Organic Nitrogen from Treated Wastewater. STAC Publication 07-001. Available online:
www.chesapeake.org.
Pagilla, K. (2007) Presentation at Water Environment Research Foundation and Chesapeake Bay
Science and Technology Advisory Committee Workshop, Baltimore, MD, September 27/28, 2007
Pehlivanoglu-Mantas, E. and D.L. Sedlak. 2006. Wastewater-Derived Dissolved Organic Nitrogen:
Analytical Methods, Characterization, and Effects - A Review. Critical Reviews in Environmental
Science and Technology. 36:261-285.
Phillips, D. 2008. Manager's Perspective on Separate Stage Suspended Growth N and DN Processes for
LOT (River Oaks). Presented at WEFTEC Technical Education Workshop No. 101, Demonstrated
Processes for Limit of Technology Nutrient Removal: Achievable Limits and Statistical Reliability.
WEF/WERF. Chicago, IL. October, 2008.
Sattayatewa, C. and K. Pagilla. 2008. Nitrogen Species Measurement in Low Total Nitrogen Effluents.
Presented at WEFTEC 2008.
Sedlak, D. 2007. The Chemistry of Organic Nitrogen in Wastewater Effluent: What It Is, What It Was,
and What it Shall Be. Presentation at the STAC-WERF Workshop: Establishing a Research Agenda for
Assessing the Bioavailability of Wastewater-Derived Organic Nitrogen in Treatment Systems and
Receiving Waters. Baltimore, MD, September, 28, 2007.
Nutrient Control Design Manual 7-21 August 2010
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Sedlak, D.L. and E. Pehlivanoglu. (2007) rDON fate and availability to nitrogen-limited algae.
Presentation at Water Environment Research Foundation and Chesapeake Bay Science and
Technology Advisory Committee Workshop, Baltimore, MD, September 27/28, 2007
Selock, K. 2008. Manager's Perspective of a Combined Process (5-Stage Bardenpho) Process for LOW N
- Severe Climate (Parkway WWTP). Presented at WEFTEC Technical Education Workshop No. 101,
Demonstrated Processes for Limit of Technology Nutrient Removal: Achievable Limits and Statistical
Reliability. WEF/WERF. Chicago, IL October, 2008.
Spani, C. 2008. Manager's Perspective on Multiple Step Chemical Addition (Rock Creek) for Low P.
Presented at WEFTEC Technical Education Workshop No. 101, Demonstrated Processes for Limit of
Technology Nutrient Removal: Achievable Limits and Statistical Reliability. WEF/WERF. Chicago, IL.
October, 2008.
STAC-WERF. 2007. Workshop Considerations and Presentations. Establishing a Research Agenda for
Assessing the Bioavailability of Wastewater-Derived Organic Nitrogen in Treatment Systems and
Receiving Waters, Baltimore, MD, September, 28, 2007.
Takacs, I. 2009. Modeling 101, How Do Influent Characteristics Affect Model Outputs? Presented at
Modeling 101 - How to Use Simulators in the Design and Operation of Wastewater Treatment
Facilities. WEF ® Distance Learning Webcast Series. Modeling Expert Group of the Americas. February
25, 2009.
Tchobanoglous, G., F. L. Burton, and H.D. Stensel. 2003. Wastewater Engineering: Treatment and
Reuse. New York, NY: McGraw-Hill.
Ten States Standards. 2004. Recommended Standards for Water Works, 2007 Edition: Policies for the
Review and Approval of Plans and Specifications for Public Water Supplies. Water Supply Committee
of the Great Lakes - Upper Mississippi River Board of State and Provincial Public Health and
Environmental Managers. Available online: http://10statesstandards.com/waterstandards.html
USEPA. 2002. Onsite Wastewater Treatment Systems Manual. Office of Water and Office of Research
and Development. EPA/625/R-00/008. Available online:
http://www.epa.gov/nrmrl/pubs/625r00008/html/625R00008.htm
USEPA. 2008a. Municipal Nutrient Removal Technologies Reference Document. Office of Wastewater
Management, Municipal Support Division. EPA 832-R-08-006. Available online:
http://www.epa.gov/OWM/mtb/publications.htm
USEPA. 2008b. Ensuring a Sustainable Future: An Energy Management Guidebook for Wastewater and
Water Utilities. January. Available online:
http://www.epa.gov/waterinfrastructure/pdfs/guidebook si energymanagement.pdf
USEPA. 2009. Accessing Individual NPDES Permits and Fact Sheets through Envirofacts. National
Pollutant Discharge Elimination System (NPDES). Website updated 13 February 2009. Website
Accessed 29 April 2009. Available online:
http://cfpub.epa.gov/npdes/permitissuance/permitscanning.cfm
Nutrient Control Design Manual 7-22 August 2010
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VA DEQ. 2004. Chapter 790. Sewage Collection and Treatment Regulations. Commonwealth of
Virginia, State Water Control Board. 9 VAC 25-790. Available online:
http://www.deq.state.va.us/export/sites/default/wastewater/pdf/DEQ SCAT Reg.pdf
WEF. 2005. Upgrading and Retrofitting Water and Wastewater Treatment Plants. Water Environment
Federation Manual of Practice, No. 28. Alexandria, VA: WEFPress.
WEF. 2009. Running An Energy-Efficient Wastewater Utility - Modifications That Can Improve Your
Bottom Line. WEF® Distance Learning Webcast Series. Moderator: J.C. Cantwell. Webcast presented June
24, 3009.
WEF and ASCE. 1998. Design of Municipal Wastewater Treatment Plants - MOP 8, 4th Ed. Water
Environment Federation and American Society of Civil Engineers. Alexandria, VA: WEF.
WERF. 2008. Dissolved Organic Nitrogen (DON) in Biological Nutrient Removal Wastewater Treatment
Processes. H.D. Stensel, Lead Editor. WERF Dissolved Organic Nitrogen Compendium. Available
online:
http://www.werf.org/AM/Template.cfm?Section=Nutrients&CONTENTID=8726&TEMPLATE=/CM/Con
tentDisplay.cfm
Eleuterio, L. and J.B. Neethling. 2009. Low Phosphorus Analytical Measurement Study. Presented at
Nutrient Removal 2009. Washington, DC. WEF.
Wisconsin DNR. 2006. Guidance for Wastewater Treatment Facility, Design Flow Determinations.
Available online : http://dnr.wi.gov/org/water/wm/glwsp/facilities/designflow.htm. Website updated
28 August 2006. Website Accessed : 29 April 2009.
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8. Selecting Candidate Treatment Processes for Plant Upgrades
Chapter 8 covers:
8.1 Introduction
8.2 Technology Selection Factors
8.3 Advantages and Disadvantages of Technology Types
8.4 Overview of Recommended Approach
8.5 Recommended Use of Advanced Tools
8.6 Patent Issues
8.7 References
8.1 Introduction
Selection of candidate treatment processes for plant upgrades is based on many factors
including target effluent limits, existing treatment, space available, and operator preference. Selection
is more of an art than a science and is influenced by the experiences and preferences of the owner and
design engineers. Guidance in Chapters 3 through 7 of this manual provides the basis for technology
selection. This chapter summarizes the key factors in selecting candidate treatment processes for
removing nitrogen, phosphorus, or both from wastewater.
8.2 Technology Selection Factors
First and foremost, the technology selected for the wastewater treatment plant (WWTP) needs
to meet the target effluent limits for nitrogen and phosphorus. See Chapter 7 for guidelines on
determining target effluent limits. Chapter 6 provides a discussion of attainable effluent limits for the
various nutrient removal technologies. It is noted in Chapter 6 and worth reiterating here that these
guidelines are not universal and that attainable nutrient removal is site specific and depends on a
number of factors including wastewater characteristics, existing plant configuration, redundancy and
reliability of equipment, and operation. This section provides a general discussion of some important
factors affecting technology selection.
8.2.1 Seasonal Permit Limits
A seasonal permit limit can have a significant effect on the design criteria and technology
selection. As noted in Chapter 7, it is common in arid regions of the country for regulatory agencies to
require nitrogen removal only during the warm months, such as from May 1 through October 31. Under
this scenario, WWTPs can stop nitrification each fall and restart it in the month of April when
temperatures are warmer to bring the plant into compliance by the beginning of May. Another
common approach is to base permit limits on annual average values. This approach allows the plant to
maximize nitrogen removal during warm months and practice variable removal during the cold months
as long as it can meet the annual average value.
Because complete nitrification can be achieved in warmer months using relatively short solids
residence times (SRTs), the volumes of biological reactors and secondary clarifiers can be significantly
reduced if the plant has a seasonal or annual average permit. Conventional biological nitrogen removal
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schemes such as the Modified Ludzack Ettinger (MLE), Anaerobic/Anoxic/Oxic 3 Stage Pho-redox (A2/0),
and Virginia Initiative Plant (VIP) processes, rather than advanced removal schemes such as the 4- or 5-
stage Bardenpho, hybrid, or fixed film systems, may provide sufficient treatment to meet the seasonal
permit conditions. Maintaining lower SRTs is not without operational disadvantages, however, because
it increases the amount of waste biomass sludge produced and therefore increases sludge processing
and disposal costs. On the positive side, reducing the SRT reduces the amount of energy required for
aeration.
There are some advantages to operating under an annual average permit rather than a seasonal
permit. It is more difficult to re-establish nitrification in April after it has been completely stopped,
which is typical of seasonal permits, than when it has been partially maintained, as is typical of annual
average permits. Also, with annual average permits, a specific effluent concentration does not have to
be met in May.
It is important to note that biological phosphorus removal systems become more efficient as
temperature decreases as long as VFAs do not become limiting and the recycle of electron acceptors
(dissolved oxygen and nitrates) to the anaerobic zone is not significant. This is because GAOs, the
primary competitors with PAOs for VFAs, depend upon glycolysis, which is more temperature sensitive
than the energy generating system of the PAOs. Therefore, as the temperature decreases below 20 °C,
the PAOs obtain an increasing fraction of the VFAs in the anaerobic zone, and their population increases
relative to the GAOs. Therefore, even though PAO biochemical reactions slow down with temperature,
as is typical of all biochemical reactors, performance often improves because of the increase in the PAO
population relative to the GAO population.
Permits may require total nitrogen (TN) removal during some parts of the year, but limit only
the ammonia to total Kjeldahl nitrogen (TKN) portion at other times of the year. If seasonal
denitrification is attempted, the acclimation period required for the biomass with some external carbon
sources needs to be accounted for in the operational start-up of denitrification. Providing sufficient time
for acclimation is important to assuring that the system is operating for removal when the limits are
applied. In some cases, operators using methanol as the carbon source for denitrification continue to
feed small doses even when the carbon is not required just to maintain the methanol metabolizing
biomass.
8.2.2 Footprint
The space available at the plant for upgrades can be the driving factor for selecting amongst
various types of nutrient removal technologies. Technologies requiring large footprints may not be
feasible in an urban area with limited space available for expansion.
Conventional nitrogen removal technologies typically have large footprints. If space is limited,
there are many alternatives for reducing basin requirements for nitrification and denitrification as
outlined in Section 8.4.
The footprint for the additional anaerobic zone for biological phosphorus removal is relatively
small, varying from 5 to 15 percent of the total volume. Space may be needed for recycle piping but the
piping frequently can be placed inside the existing reactor (which enables the use of low head propeller
pumps). In many retrofits, the anaerobic zone can be created in an existing activated sludge basin using
baffles.
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The footprint for adding chemical feed facilities to precipitate phosphorus is typically small. A
new solids removal process (tertiary clarification, effluent filtration) may require a larger footprint. High-
rate clarification (HRC) can be used instead of conventional clarification to reduce the footprint size.
Examples of patented HRC systems are provided in Chapter 6.
8.2.3 Hydraulic Considerations
Certain processes require additional pressure head, which is supplied by pumps. For example,
pumping is typically required for plants with denitrifying filters because sufficient hydraulic head
following the aeration basins is not usually available. Pumping is also usually required for effluent
filtration. The addition of overflow baffles with different elevations to a treatment system may induce
some head losses and require additional pumping. Underflow baffles provide an equal amount of
protection against backflows if correctly designed, and do so without headless.
Membrane bioreactors require additional pumping to force water through the membrane
material. Pressure-driven systems are used if the membrane is configured after the aerobic and anoxic
basins. Vacuum-driven systems use a pump to pull water through the membrane when it is submerged
in a biological reactor.
8.2.4 Chemical Needs
The addition of any commercial chemical requires delivery, storage, and safety procedures in
addition to mixing and dosing equipment. Three types of chemicals may be added to wastewater for
nutrient removal: metal salts or lime to precipitate phosphorus; an external carbon source for BPR or
denitrification; and lime, bicarbonates, or caustic soda to maintain alkalinity.
Metal salts or lime can be added to wastewater to precipitate phosphate and subsequently
remove it using a solids separation process. Many plants that are required to meet low effluent total
phosphorus (TP) limits that use BPR are incorporating chemical phosphorus removal as a back-up system
to ensure that limits can be met in the event of a biological upset. Multiple feed points are also
becoming more common.
As noted in the previous section, an external organic carbon source may be needed for BPR or
denitrification. Common external sources include methanol, ethanol, and acetic acid, as well as simple
sugars.
The nitrification process consumes alkalinity. Although the denitrification process can add
alkalinity back to the mixed liquor for single-sludge treatment systems, there is still typically a net loss in
alkalinity of 50 percent or more. Alkalinity depletion can result in pH decreases and interfere with
biological processes. As previously noted, plants can add alkalinity to maintain adequate levels by
adding lime, bicarbonates, or caustic soda to the influent.
8.2.5 Available Sludge Treatment and Options
As noted in Chapter 4 of this manual, less sludge production and better settling and thickening
sludge is found for activated sludge systems using anoxic/aerobic treatment for nitrogen removal versus
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aerobic treatment only. The impact this will have on total sludge production depends upon how much
waste sludge is produced by other treatment units such as primary clarifiers and treatment with
precipitating chemicals. As noted in Chapter 5, sludge from BPR processes will have higher phosphorus
contents, and therefore a higher settling velocity, but will otherwise be similar to sludge from
conventional activated sludge plants.
Additional sludge production is seen as one of the major disadvantages of chemical phosphorus
removal compared with biological methods. The volume of additional sludge produced varies
depending on the influent phosphorus concentration, chemical application point(s), chemical used, and
target effluent limit. Chemically treated sludge has a higher inorganic content compared to primary and
activated sludge and will increase the required size of aerobic and anaerobic digesters, as well as
dewatering equipment. The use of metal salts can result in increased inorganic salts (salinity) in the
sludge and in the effluent. Salinity can create problems when biosolids are land applied or when the
effluent is returned to existing water supply reservoirs. Lime traditionally produces a higher sludge
volume compared to metal salts because of its reaction with natural alkalinity.
Designers should carefully consider the impacts of sludge processing on nutrient release into
recycle streams. Anaerobic digestion can release ammonia and phosphorus in large quantities, resulting
in significant recycle loads to the biological nutrient removal (BNR) processes. Anaerobic digestion of
BPR sludges also releases magnesium along with the ammonia and phosphorus, resulting in the
formation of struvite. This struvite typically remains entrapped in the digested solids and reduces the
loads in the liquid stream recycled back to the mainstream processes. Although phosphorus removal by
struvite formation is a desirable phenomenon from the perspective of recycle load, struvite crystals can
plug centrifuge ports, as well as pumps and pipes used to convey the sludge, if not controlled. However,
because of the higher concentrations of magnesium and phosphorus in anaerobic digesters when BPR
sludges are digested, most, if not all, of the struvite formation occurs within the sludge and remains with
the sludge rather than staying in solution and forming struvite after leaving the digesters.
Innovative work is underway to recover nutrients from sludge (including struvite) and convert
them into commercial fertilizers and other products in which they can be reused. See Chapter 14 for
discussion of nutrient recovery technologies.
8.2.6 Energy Considerations
Treatment plant upgrades for nutrient removal often require additional energy to operate. If
they do not already exist, addition of recycle streams for BNR will require additional pumping energy.
Filters likely require additional electricity for pumping to provide the pressure head on the filters.
Membranes have the highest pumping cost because they require the most pressure to push the water
through the membrane material. The smaller the membrane pore size, the higher the pumping costs.
Anaerobic zones, fermenters, and chemical feed pumps usually require a relatively small amount of
energy (USEPA 2008a). Although some processes are inherently less energy intensive than others (e.g.,
oxidation ditch vs nitrifying diffused air activated sludge), energy requirements generally increase with
level of treatment. The exception is when a primary anoxic zone is added for denitrification when
nitrification is already being practiced. The reduction of biochemical oxygen demand (BOD) in the
anoxic zone reduces the amount of energy needed to transfer oxygen in the aerobic zone. If the BNR
system is an MLE or an A2/O configuration, the anticipated reduction in aeration energy is commonly
about 20 percent. Also, using low head propeller pumps, or equivalent, and nitrate recycle lines that are
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internal to the activated sludge basins will significantly reduce pumping energy compared to recycle
pump stations.
Energy requirements for nutrient removal can be further minimized by implementing the
following (Kang et al. 2009):
• Use of swing zones
• Step aeration in activated sludge basins
• Fermenters for in-plant carbon generation
• Automated control of aeration equipment
• Sidestream treatment
• Flow equalization for recycle flows
Energy required for nutrient removal can also be offset by energy conservation measures
(ECMs). ECMs range from no-cost operational improvements such as peak load shaving and
optimization of mixing and aeration, to plant upgrades such as pump replacements for more efficient
variable frequency drives. Wastewater utilities have an opportunity to be energy generators by using
biogas from anaerobic digestion to generate electricity onsite. Many energy efficiency measures have a
short payback period for the wastewater utility to realize positive cost benefits.
Implementation of energy efficiency measures begins with benchmarking current performance
and conducting an energy audit to identify areas of improvement. EPA has developed guidelines for
energy efficiency based on the Plan-Do-Check-Act management system approach to help operators
identify, implement, measure, and improve energy efficiency and renewable opportunities at their
utilities. For more information and guidance, refer to the document, Ensuring a Sustainable Future: An
Energy Management Guidebook for Wastewater and Water Utilities, available online at
http://www.epa.gov/waterinfrastructure/pdfs/guidebook si energymanagement.pdf.
8.2.7 Staffing and Training Requirements
A potential disadvantage to some emerging technologies is the amount of training required to
operate a non-conventional system. Even though emerging technologies, over the long term, may be
less complex than conventional systems, operators may have a preference for a more conventional type
system that fits in with the existing technologies. However, many operators like to learn about
innovative processes and frequently learn to operate them in ways that overcome design limitations.
BNR processes have been operated very well by many operators after a relatively short training period.
It is especially important that the operators understand the major operating factors that affect the
performance of the BNR system. Automated process control can ease the burden on operators and is
recommended for systems targeting very low effluent TN and TP levels.
8.2.8 Technology Selection Considerations for Small Flow Systems
Nutrient removal may be required for very small or decentralized wastewater treatment
applications. Unique features of these treatment applications with regard to wastewater
characteristics, operations and maintenance capabilities, and water reuse potential affect the process
selection and design approaches. In contrast to larger flow systems, the diurnal variations in the
wastewater flow rates and concentrations are much more dramatic, with a large increase in flow and
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wastewater strength after day break and with minimal wastewater flows in the late evening and early
morning hours. The wastewater strength is also typically much greater than that found for centralized
wastewater treatment facilities much of the time, which greatly affects the process selection
alternatives to meet low effluent nutrient concentrations.
Small flow and decentralized wastewater treatment systems generally have less resources for
operational and maintenance staff, which favors selecting processes that have less mechanical
complexity and/or exhibit a high degree of reliability, and process designs with higher safety factors.
Many small flow and decentralized wastewater treatment facilities are located near applications for
water reuse, such that the process selection includes both nutrient removal and required water quality
for water reuse. As the need for higher treatment levels has evolved for small systems, many of the
processes commonly used in the past may be outdated and new approaches are continually being
developed. For additional guidance for small flow systems, please consult the EPA website at:
http://cfpub.epa.gov/owm/septic/septic.cfm7page id=268.
8.3 Advantages and Disadvantages of Technology Types
To assist in technology selection, Table 8-1 provides some very general advantages and
disadvantages of some common technology types. A discussion follows the table.
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Table 8-1. Advantages and Disadvantages of Technology Types
Technology
Type
Advantages
Disadvantages
Attached growth
compared to suspended
growth
Smaller footprint
Can biologically nitrify at
lower SRTs
More resilient to shock loads
Less biological sludge
production
Better settling sludge
High oxygen transfer requirements
Operation is less flexible (i.e.,
accumulation of nitrifying biomass
is limited because of competition
with heterotrophic microorganisms
for DO and carbon)
Secondary process such as
chemical precipitation is needed
for phosphorus removal
Chemical phosphorus
removal compared to
biological phosphorus
removal
Not susceptible to biological
upsets and is, thus, generally
considered more reliable and
controllable.
Can increase sludge settling
rates.
Increased sludge volume
Consumes alkalinity, which can
affect nitrification
If added to secondary treatment,
increased MLSS concentration and
therefore, increased clarifier
volume
Chemical costs
Mainstream treatment
compared to sidestream
treatment of recycle
streams
Flow equalization can help
with wet weather flows and
can provide operational
flexibility
Mainstream treatment is
straightforward with existing
processes, whereas
sidestream treatment may
require operating training
and vendor support
Equalization (EQ) basins require
additional space, high capital costs.
Sidestream treatment requires an
additional but smaller footprint
than EQ basins
Continuous sludge dewatering
requires 24-hour operation vs.
daily operation
Emerging technologies
compared to established
technologies
• Can offer improved
efficiency and operational
cost savings
Limited experience under a variety
of operating conditions
Attached growth systems can be used to considerable advantage for nitrification and
denitrification—biological aerated filters (BAF) being an example—but they are difficult, though not
impossible, to use for biological phosphorus removal. Consequently, when BAFs are used and
phosphorus removal is also required, the phosphorus is almost always removed by chemical
precipitation. The BAFs have a small footprint, which can be a major advantage, but they tend to have
higher oxygen transfer energy costs when used for nitrification, and organic carbon has to be added for
denitrification. Flexibility for treatment of changing flows and loads only comes from having multiple
units, some of which can be taken on and off line at the appropriate times.
The addition of fixed film media to an activated sludge basin has several advantages. The
attached biomass does not flow forward to the secondary clarifier, reducing the solids loading to the
clarifier. The SRT of the attached biomass is long, so that the average SRT of the attached and
suspended biomass is long enough to accomplish nitrification under conditions that would not be
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conducive to nitrification with suspended biomass alone. The disadvantages are the costs and
installation of the media and necessary appurtenances such as screens, air knifes, recycle pumps and
mixers, depending upon the type of media selected.
Chemical phosphorus removal using iron or aluminum salts can be used instead of BPR or as a
supplement to BPR. The advantages of chemical phosphorus removal are that (1) the chemical addition
can be turned on and off without an acclimation period or harm to the biological processes, (2) removals
can be obtained even when there are biological upsets, (3) the process is easy to control for consistent
removal to low levels, and (4) the addition of the chemicals increases the sludge settling rates when
added directly to the mixed liquor. Disadvantages of chemical addition are (1) the increase in sludge
production; (2) the cost of the chemicals and consumption of alkalinity by both iron and aluminum
chemicals, which may result in a need to add alkalinity for both chemical coagulation and nitrification;
(3) and increases in the MLSS concentrations in the reactors. It is important to note that addition of
VFAs to a biological phosphorus removal process makes it easier to control similar to chemical addition
for phosphorus removal, with a lesser increase in waste sludge production.
Sidestreams can be managed at the plant through mainstream treatment or can be treated
separately using a conventional or advanced sidestream treatment process. Consistently low effluent
TN does not require removal of nitrogen from sidestreams; mainstream treatment such as flow
equalization or continuous dewatering operations can be just as effective (Barnard 2006). Mainstream
treatment can be a lower cost option but often requires additional monitoring and controlled operation.
Also, operators are more familiar with mainstream as compared to advanced treatment options.
Advanced sidestream treatment methods, however, have been increasingly used in recent years. They
typically have smaller overall footprints. See Chapter 6, Section 6.6 for a description of several advanced
sidestream treatment processes.
8.4 Overview of Recommended Approach
Although selection of the most appropriate nutrient removal technology depends on many
factors, designers can begin with a generalized approach based on a few important external conditions,
namely:
• Target effluent limits for TN
• Target effluent limits for TP
• Available Space
• Temperature
As noted throughout this design manual, simulators are the best tool for process selection and design.
See Chapter 6 for detailed information on the available technologies for nitrogen and phosphorus
removal at WWTPs and Chapter 10 for guidance on process selection and design using models and
simulators.
1. Target effluent limits for TN
• For up to 70 percent removal of TN from the wastewater and typically producing effluent
concentrations between Sand8mg/L, depending upon the internal recycle rate, the basic
configuration of a single anoxic and aeration zone (e.g., the MLE process) with conventional
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secondary clarifiers is often sufficient. Target effluent levels between 5 and 8 mg/L year
round, however, may be difficult to achieve with this technology in areas with very cold
climates or where water conservation has led to a more concentrated wastewater.
Typically, however, effluent concentrations in that range can be obtained if the internal
recycle rate can be increased to higher rates. This assumes the design permits the RAS rate
to be increased to at least 100 percent of the influent, and preferably to 150 percent.
• For 70 to 90 percent removal of TN from wastewater and typically producing effluent
concentrations between 3 and 5 mg/L, a second anoxic zone can be added after the aerobic
zone such as in the 4-stage Bardenpho process configuration, or an effluent denitrification
filter can be used to achieve the required removal by denitrifying the portion of the flow
that is not recycled to the first anoxic zone. Complete denitrification by the 4-stage
Bardenpho process typically requires addition of an external carbon source to the second
anoxic zone unless a very large second anoxic zone is used. Removal can be improved
through optimization of influent loading by equalization, or by controlling recycle streams
and influent flow to prevent washout of nitrifiers.
• For 90 percent or more removal of TN from wastewater and typically producing effluent
concentrations of 3 mg/L or less, removal may become limited by dissolved organic
nitrogen (DON) in the plant effluent. Based on a study of 188 plants, effluent DON (or
EDON) typically ranges between 0.5 and 1.5 mg/L (Pagilla 2007). Research is ongoing into
the design and operating conditions that can be used to minimize EDON. Process
optimization and automated control systems become very important in achieving these low
TN levels. Denitrification filters with supplemental carbon (e.g., methanol) addition can
provide enhanced TN removal consistently to low effluent levels.
• See Chapter 6, Section 6.7, for additional discussion of attainable effluent TN.
2. Target effluent limits for TP
• For 80 to 90 percent removal of TP from wastewater and typically producing effluent
concentrations of 0.5 to 1.0 mg/L, BPR should be sufficient as long as the following
conditions are met:
- The ratio of readily biodegradable carbonaceous oxygen demand (rbCOD) to TP
(rbCOD:TP) is sufficient. Typically, BPR will work for rbCOD:TP ratios of 10 to 16 or
greater; see Chapter 5, Section 5.2 of this design manual for additional information.
- The recycle of dissolved oxygen and nitrates to the anaerobic zone are controlled.
- Large amounts of phosphorus are not released during sludge handling and recycled to
the biological process influent.
To consistently achieve low levels under all conditions, chemical addition should be
considered as a backup.
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• For 90 percent or more removal of TP from wastewater and typically producing effluent
concentrations less than 0.5 mg/L TP, it is common for BPR plants with primary settling to
need supplemental chemical addition unless sludge fermentation or supplemental organic
carbon addition is practiced. The key to achieving low TP limits is the solids separation
process. Conventional clarification is not sufficient in many cases. Low target limits may
require advanced clarification processes or effluent filtration. Note that effluent TP
concentrations of 0.2 mg/L or even less can be consistently achieved with BPR if adequate
amounts of VFAs are present in the anaerobic zone (through favorable rbCOD:TP ratios in
the influent or through VFA addition) and recycle streams are carefully controlled.
• See Chapter 6, Section 6.7 for additional discussion of attainable effluent TP.
3. Available Space. Retrofitting for BPR or adding chemicals for phosphorus removal requires very
little in the way of additional space at the treatment plant. The footprint for the anaerobic zone
for BPR is typically 5 to 15 percent of the total anaerobic/aerobic volume, and this can often be
created using baffles in the existing basin. Chemical dosing equipment is relatively small and
can usually be configured within the existing plant layout. Advanced solids/liquid separation
technologies such as tertiary clarification or filtration can, however, have large footprints and
require additional space. High-rate ballasted clarification (HRBC) should be considered when
space is constrained.
Although space requirements for phosphorus removal are typically low, suspended growth
aeration processes for nitrogen removal require high SRTs and thus, large basins compared to
conventional activated sludge processes. If sufficient volume is available in the existing
activated sludge basin, it can be modified to configure separate anoxic and aerobic zones by
adding baffles, mixers and an internal recycle. The configuration of the activated sludge basin
determines the complexity of the internal recycle system. Sometimes all that is needed is a port
in one of the walls and a propeller pump installed in the port to pump the recycle through the
wall to the adjacent basin. If sufficient volume is not available in the existing activated sludge
basins (which is often the case), additional basins may be needed. Designers should consider
the following approaches to minimize the footprint for biological nitrogen removal when space
is constrained:
• Increase the biomass in the activated sludge basins using a combination suspended-
attached growth system such as Integrated Fixed Film Activated Sludge (IFAS). This will
reduce the size of the biological reactors and, therefore, the footprint.
• Use membrane bioreactors (MBR) for liquid-solids separation instead of conventional
clarification. MBR systems can operate at a higher mixed liquor suspended solids (MLSS)
concentration, thereby reducing basin requirements for the biological reactors.
• Use an attached growth reactor for nitrification such as a Biological Aerated Filter (BAF).
• Instead of a second anoxic zone, consider a denitrification filter. Denitrification filters have
the added advantage of achieving low effluent TN limits but usually require a supplemental
carbon source. They also require additional pumping to provide the pressure head to move
water through the filter.
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• Consider bioaugmentation of nitrifiers to reduce the needed SRT in the activated sludge
basin and thus provide more treatment in a smaller volume.
• If peak flows are the issue, consider a step feed system to handle wet weather events
instead of constructing additional basins.
4. Temperature. Biological treatment processes are sensitive to changes in temperature. The
most sensitive process is nitrification, with the rate approximately doubling for every 8 to 10° C
increase in temperature (WEF and ASCE, 2006). It is very important that designers consider the
annual temperature range, not just the design temperature when selecting candidate
technologies. In temperate regions of the country, fixed film systems such as BAF and hybrid
systems such as IFAS compensate for temperature impacts by maintaining the nitrifiers in the
system for very long SRTs. A swing zone that allows operators to switch between anoxic and
aerobic zones also can help ensure consistent removal under varying temperature conditions.
Low temperatures can also interfere with fermentation of wastewater in sewers and anaerobic
zones and reduce the production of VFAs required for BPR. Onsite sludge fermentation has
successfully been used in colder climates to address this issue. As previously noted, BPR process
performance often improves with lower temperatures because of microbial population shifts
that result in larger populations of PAOs relative to the GAOs, assuming the recycle of electron
acceptors to the anaerobic zone is controlled. See Sections 4.5.3, 5.4.4, and 8.2.1 of this manual
for additional information on the effects of temperature on BNR.
8.5 Recommended Use of Advanced Tools
This section introduces three advanced tools that can be used for designing upgrades for
nutrient removal: pilot testing, mathematical modeling, and bench scale testing.
Pilot testing of the preferred nutrient removal alternative is very common in the wastewater
industry. Testing typically takes 6 months to 1 year depending on seasonal variability and can be done
during the planning or preliminary design phase. Although it can be expensive, pilot testing allows
designers to size and configure unit processes based on actual wastewater and biomass characteristics
and can reduce safety factors and thus, save capital costs considerably in excess of the cost of the study.
Mathematical equations representing wastewater treatment processes and simulation
programs used to solve them have become increasing powerful and common in recent years. Computer
modeling has many benefits over traditional design approaches. Designers can use models to quickly
test many different treatment configurations and design parameters in a fraction of the time and cost of
performing traditional design calculations and laboratory or pilot tests. While setting up and calibrating
models can be a complex process, it is ultimately easier and more powerful than traditional design
methods, when done properly. Modeling has therefore become the preferred method of design for
nutrient removal at WWTPs. Chapter 10 provides additional discussion of the benefits of mathematical
modeling along with detailed guidance on setting up and running a process simulation model for
nutrient removal.
Data from bench scale tests are commonly used to calibrate process simulation models. One of
the most important tests for nitrification is the determination of maximum specific nitrifier growth rate.
Chapter 10 discusses the importance of this parameter and summarizes proven characterization
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methods. Bench scale tests can also provide very useful information on the relative fractions of
particulate, soluble, and readily biodegradable organic material (which is very important for biological
phosphorus removal) and can be used to estimate rDON. Determination of non-biodegradable
suspended solids fractions is also very important for accurate estimates of design MLSS concentrations.
Because reactions are complex and difficult to generalize, jar testing for chemical phosphorus removal is
also highly recommended. Bench scale testing is also common and recommended for evaluating
alternative carbon sources.
8.6 Patent issues
A patent is essentially a type of property right. It gives the patent holder the right to limit others
from making, using, offering for sale, or selling the invention in the United States or importing the
invention into the United States. In general, the term of a patent is 20 years from the date that the
patent application was submitted. The U.S. Patent and Trademark Office (USPTO) is the government
agency responsible for reviewing applications and issuing patents.
There are three types of patents (USPTO 2009):
• Utility patents may be granted to anyone who invents or discovers any new and useful process,
machine, article of manufacture, or composition of matter, or any new and useful improvement
thereof.
• Design patents may be granted to anyone who invents a new, original, and ornamental design
for an article of manufacture.
• Plant patents may be granted to anyone who invents or discovers and reproduces any distinct
and new variety of plant.
Patents for wastewater treatment technologies are usually utility patents.
If a patented process is selected for nutrient removal, license and royalty fees may apply and
should be included in the cost analysis. Fees are usually negotiated between the patent holder and the
plant owner. To search for patents, use the USTPO database, available online at http://patft.uspto.gov/
or visit a patent and trademark depository library. Refer to
http://www.uspto.gov/go/ptdl/ptdlib l.html for library locations in your state.
In general, patent and licensing fees are not required for biological phosphorus removal
systems. Some of the biological phosphorus removal technologies listed in Chapter 6 of this manual
may have at one time been patented, but many have been in use longer than the patent period.
Proprietary systems for attached growth nitrification and sidestream processes are patented, but the
patent fees are included in the price from the manufacturer.
Nutrient Control Design Manual 8-12 August 2010
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8.7 References
Barnard, J.L 2006. Biological Nutrient Removal: Where We Have Been, Where We are Going? In
Proceedings of the Water Environment Federation, WEFTEC 2006.
Bilyk, K., J. Rohrbacher, T. Bruton, P. Pitt, R. Latimer. 2009. Evaluating the Cost and Process
Performance of Carbon Alternatives to Methanol. Presented at Nutrient Removal 2009. Washington,
DC. WEF.
deBarbadillo, C, J. Barnard, S. Tarallo, and M. Steichen. 2008. Got Carbon? Widespread biological
nutrient removal is increasing the demand for supplemental sources. Water Environment &
Technology. Alexandria, VA: WEF. 20(1): 49-53.
deBarbadillo, C., P. Miller, and S. Ledwell. 2009. A Comparison of Operating Issues and Dosing
Requirements for Alternative Carbon Sources in Denitrification Filters. Presented at Nutrient Removal
2009. Washington, DC. WEF.
Copithorn, R. R., G. D. Farren, D. Sen and C. W. Randall. 1993. Full-Scale Evaluation of Nitrification and
Denitrification on Fixed-Film Media (Ringlace) for Design of a Single-Sludge Nitrogen Removal System,
Proceedings Liquid Treatment Processes, 66th Annual Conference Water Environment Federation,
(3):137-148. Anaheim, California, October 30, 1993.
Cherchi, C., A. Onnis-Hayden, and A.Z. Gu. 2009. Ability of Specific Enriched Denitrifying Culture to
Utilize Other Carbon Sources. Presented at Nutrient Removal 2009. Washington, DC. WEF.
Gellner, W. J., L Samel, D. Howard, Al. Stone, and P. Pitt. 2008. When Conventional Design Bid Build
Approach is Unconventional Design and Bidding of IFAS Full-Scale Pilot at Greensboro, North Carolina.
In Proceedings of the Water Environment Federation, WEFTEC 2008.
Johnson, T. L, J. McQuarrie, and A. Shaw. 2004. Integrated Fixed-Film Activated Sludge (IFAS): The
New Choice for Nitrogen Removal Upgrades in the United States. In Proceedings of the Water
Environment Federation, WEFTEC 2004.
Kang, S.J., K.P. Olmstead, K.M. Takacs, J. Collins, J. Wheeler, and P. Zharaddine. 2009. Sustainability of
Full-Scale Nutrient Removal Technologies. Presented at Nutrient Removal 2009. Washington, DC.
WEF.
Maquarrie, J., D. Waltrip, R. Rutherford, W. Thomas, C. Bott, R. Baumler, and D. Katehis. 2009. Full-
Scale Design Challenges for the James River Treatment Plant IFAS Improvement Project: Using a
Demonstration Scale Project o Optimize Final Design. Presented at Nutrient Removal 2009.
Washington, DC. WEF.
Majed, N., A. Onnis-Hayden, T. Welander, and A.Z. Gu. 2009. Decoupling and Optimization of Both P
and N Removal in an Advanced IFAS-EBRP-MBR System. Presented at Nutrient Removal 2009.
Washington, DC. WEF.
Onnis-Hayden, A., N. Majed, K. McMahon., A. Gu. 2008. Phosphorus Removal and PAOs Populations at
a Full-Scale Integrated Fixed Film Activated Sludge (IFAS) Plant. In Proceedings of the Water
Environment Federation, WEFTEC 2008.
Nutrient Control Design Manual 8-13 August 2010
-------
Phillips, H. M. and E. Kobylinksi. 2007. Sidestream Treatment vs. Mainstream Treatment of Nutrient
Returns when Aiming for Low Effluent Nitrogen and Phosphorus. In Proceedings of the Water
Environment Federation, WEFTEC 2007.
Randall, C. W., and D. Sen. 1995. Full-Scale Evaluation of an Integrated Fixed-Film Activated Sludge (IFAS)
Process for Enhanced Nitrogen Removal. Wat. Sci. Tech. (33)12: 155-162. Elsevier Science Ltd., GB.
Sriwiriyarat, T., and Randall, C.W. 2005. Evaluation of IFAS Wastewater Treatment Processes at High
MCRT and Low Temperatures. Journal of Environmental Engineering. ASCE. 131(11): 1550-1556.
Tchobanoglous, G., F. L. Burton, and H.D. Stensel. 2003. Wastewater Engineering: Treatment and
Reuse. New York, NY: McGraw-Hill.
Thomas, W.A., C.B. Bott, P. Regmi, G. Schafran, J. McQuarrie, B. Rutherford, R. Baumler, and D. Waltrip.
2009. Evaluation of Nitrification Kinetics for a 2.0 MGD IFAS Process Demonstration. Presented at
Nutrient Removal 2009. Washington, DC. WEF.
USEPA. 2008a. Municipal Nutrient Removal Technologies Reference Document. Office of Wastewater
Management, Municipal Support Division. EPA 832-R-08-006. Available online:
http://www.epa.gov/OWM/mtb/mnrt-volumel.pdf
USEPA. 2008b. Emerging Technologies for Wastewater Treatment and In-Plant Wet Weather
Management. EPA 832-R-06-006. Available online: http://www.epa.gov/OW-
OWM.html/mtb/emerging_technologies.pdf
WEF and ASCE. 1998. Design of Municipal Wastewater Treatment Plants - MOP 8, 4th Ed. Water
Environment Federation and American Society of Civil Engineers. Alexandria, VA: WEF.
WEF and ASCE. 2006. Biological Nutrient Removal (BNR) Operation in Wastewater Treatment Plants -
MOP 29. Water Environment Federation and the American Society of Civil Engineers. Alexandria, VA:
WEFPress.
WERF. 2008. Carbon Augmentation for Biological Nitrogen Removal. Julian Sandino, Lead Editor.
WERF Dissolved Organic Nitrogen Compendium.
U.S. Patent Office. 2009. United States Patent and Trademark Office, Main Webpage. Website
Accessed 8 May 2009. http://www.uspto.gov/main/patents.htm
Nutrient Control Design Manual 8-14 August 2010
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9. Design Approach for Phosphorus Removal by Chemical Addition
Chapter 9 covers:
9.1 Introduction
9.2 Selecting a Chemical Precipitant
9.3 Selecting Point(s) of Application
9.4 Determining Chemical Dose
9.5 Designing a Chemical Feed System
9.6 Designing for Rapid Mixing and Flocculation
9.7 Solids Separation Processes
9.8 Operational Factors
9.9 References
9.1 Introduction
Many wastewater treatment plants (WWTPs) that are required to remove phosphorus do so by
adding chemicals to precipitate phosphate and then remove it using solids separation techniques.
Chemicals may be added to primary, secondary, or tertiary processes, or at multiple locations in the
plant. Chemicals used for phosphorus precipitation include metal salts, such as ferric chloride and
aluminum sulfate (alum), and lime.
Chapter 3 explained the principles of phosphorus removal by chemical addition. The purpose of
this chapter is to provide guidance on key design and operational issues. This chapter's discussion on
solids separation focuses on advanced wastewater clarification, while effluent filtration is discussed in-
depth in Chapter 11.
9.2 Selecting a Chemical Precipitant
Chemicals for precipitation are based on either aluminum (III), ferric (III), ferrous (II), or calcium
(II) compounds that react with soluble phosphorus. The most common chemicals used are aluminum (III)
and ferric (III) compounds; in particular, the two metal salts aluminum sulfate (alum) and ferric chloride.
Alum and ferric chloride have similar properties and performance. Compared with lime, they create less
sludge and tend to be more popular with operators. Polymers can be added with metal salts to improve
settling or the removal of precipitated phosphate via filtration.
9.2.1 Advantages and Disadvantages of Metal Salts
Table 9-1 lists metal salts available for phosphorus removal and provides a summary of available
forms and storage requirements, advantages, and disadvantages. A discussion of the advantages and
disadvantages of lime is provided in section 9.2.2. Note that phosphorus removal by metal salts is based
on stoichiometry whereas removal by lime is based on achieving a target pH range.
Nutrient Control Design Manual
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Table 9-1. Advantages and Disadvantages of Common Aluminum and Iron Salts
Type
Available Forms and Storage Issues
Advantages
Disadvantages
Aluminum
Aluminum
sulfate (alum)
Sodium
aluminate1
Polyaluminum
Chloride (PACI)
• Liquid or dry form
• Stable at room temperatures in
closed containers under normal
storage conditions
• Near unlimited shelf life
• Corrosive, dust is irritant
• Liquid or dry form
• Liquid has 2-3 month shelf-life,
strong alkali, handle as caustic
• Dry has 6 month shelf-life, non-
corrosive, dust is irritant
• Available as liquid, hydrate
form. Available in different
strengths
• Corrosive mineral acid.
Requires acid-resistant
materials like PVC, Teflon®,
rubber, and ceramic materials
• Most common form of
aluminum salt. Has been
used to achieve low effluent
phosphorus concentrations
• Does not consume alkalinity
• Appropriate for low
alkalinity wastewater or
where pH is already low and
further depression should
be avoided
• Does not change pH of
wastewater
• Various formulations
(multiple aluminum atoms
bonded with chloride)
available depending on
manufacturer
• Can help lower turbidity
• May need excess to depress
the pH to an optimal
operating environment
• Alum sludge may be more
difficult to thicken and
dewaterthan iron sludge
• Dissolved C02 or other acidity
is needed to avoid pH
increase above optimum
zone
• Performance considered
inferior to alum
• Not compatible with carbon
steel, stainless steel, brass or
aluminum
• Unit cost may be higher
Iron
Ferric chloride2
Ferrous chloride
Ferrous sulfate
• Available as liquid
• Very corrosive
• Stains concrete and other
materials
• Available as liquid
• Slightly less corrosive than
ferric chloride
• Available dry
• Acidic when dissolved
• Oxidizes in moist air
• Cakes at storage temp above
20°C
• More common than ferrous
chloride or ferrous sulfate
• Has been used to achieve
low effluent phosphorus
concentrations
• Can be available as a low-
cost industrial byproduct
• Can be available as a low-
cost industrial byproduct
• Especially corrosive and
requires special piping
• In plants with poor solids
capture, ferric chloride might
impart a slight reddish color
to effluent
• Can be an issue if UV
disinfection is downstream3
• If industrial byproducts, may
have large amounts of
impurities, such as free acid
or metals
• Produces low phosphorus
levels only at high pH
• If industrial byproducts, may
have large amounts of
impurities, such as free acid
• Produces low phosphorus
levels only at high pH
1. Granular trihydrate is a common commercial form
2. "Ferric" is also a common trade name for FeCI3 and also 40 percent liquid FeCL3 solution.
3. Potential for problems depends on chemical application point. Of great concern if it is added in chemical treatment
process preceeding UV disinfection.
Source: WEF and ASCE (1998); USEPA (1987a); Gulbrandsen (2008); Tchobanoglous et al. (2003).
Nutrient Control Design Manual
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9.2.2 Advantages and Disadvantages of Lime
As explained in Chapter 3, lime first reacts with bicarbonate alkalinity to form calcium
carbonate. As the pH increases, excess calcium ions will react with the phosphorus to precipitate
hydroxylapatite (Ca5(OH)(PO4)3).
The primary reasons for infrequent use of lime for phosphorus removal are:
• The substantial increase in the mass of sludge to be handled compared to that from use of
metal salts.
• Phosphorus precipitation to low concentrations occurs at high pH levels (11-12), and pH
neutralization is required before effluent discharge.
• The operation and maintenance problems associated with the handling, storage, and
feeding of lime.
Lime sludge can reach 0.5 percent of the volume of wastewater treated (Tchobanoglous et al.
2003). To avoid this large production of sludge, plants can consider adding lime in a tertiary process.
Tertiary designs can be single-stage with one rapid mix, flocculation, sedimentation, and recarbonation
basin or a two-stage process with a second sedimentation basin, which allows for better control and
higher removal. Addition of lime to primary tanks has the advantage, however, of removing substantial
amounts of total suspended solids (TSS) and biochemical oxygen demand (BOD), reducing secondary
treatment requirements.
The key variable in phosphorus removal using lime is alkalinity. If the alkalinity is too low, lime
addition will create a poorly settleable floe (WEF and ASCE 1998). If the alkalinity is moderate to high (>
150 milligrams per liter (mg/L)), the floe will contain a greater fraction of dense calcium carbonate
precipitate, which leads to enhanced settling. WEF and ASCE (1998) report excellent phosphorus
removal with high alkalinity waters with lime treatment to a pH of 9.5 to 10, but higher pH values are
commonly used. Regardless of the application point, pH adjustment is needed following lime treatment
and is usually accomplished by adding carbon dioxide or a liquid acid such as sulfuric acid, nitric acid, or
hypochlorite (Tchobanoglous et al. 2003; USEPA 1999).
Chemical costs for lime can be reduced by recovering lime on-site using lime recalcination.
Recalcination is achieved by heating the sludge to 980 °C to convert calcium carbonate to lime. The
carbon dioxide from this process can be used as a source for recarbonation of the water for pH
adjustment.
Lime requires special handling that sets it apart from metal salts. Carbonate scaling can form on
equipment and pipes. Lime slaking, where quicklime (CaO) is reacted with water to form Ca(OH)2, is the
biggest operational disadvantage. Section 9.5.2.3 provides a discussion of lime slaking.
Although lime had lost favor due to issues associated with chemical handling and sludge
production, it has been regaining popularity because of its ability to reduce phosphorus to very low
levels when combined with effluent filtration, as well as its beneficial microbial control properties during
sludge digestion. WEF and ASCE (1998) report that residual phosphorus concentrations of 0.1 to 0.2 can
be achieved with lime treatment and granular media filtration. The Upper Occoquan Sewage Authority
Advanced Wastewater Treatment Plant, Fairfax County, VA, has routinely and consistently achieved
effluent soluble phosphorus concentrations of less than 0.05 mg/L using high lime treatment followed
Nutrient Control Design Manual 9-3 August 2010
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by sand filtration since July of 1978, a period of more than thirty years. The effluent requirement for the
facility over the entire period of time has always been 0.1 mg/L TP.
9.2.3 Costs
According to the U.S. Geological Survey (USGS) 2007 Minerals Yearbook, the cost for quicklime
in 2007 was $76.70 per ton, and hydrated lime cost $92.90 per ton (Miller 2008). Dry alum can cost
around $260 per ton, and a ton of liquid ferric chloride can cost around $400 (Yuba City 2007; City of
East Moline 2007; City of Santa Cruz 2006; Saginaw City Council 2006). Lime and other chemical costs
are expected to continue to rise in the coming years with increasing global demand for chemicals, rising
oil prices, and potential supply shortages, which can be worsened by certain local conditions like
weather events that disrupt distribution systems (Plank 2008).
9.3 Selecting Point(s) of Application
Metal salts can be added at one or multiple locations within a treatment plant to remove
phosphorus (see Figure 9-1 for possible points of chemical addition for a conventional WWTP). "Pre-
precipitation" is when chemicals are added to raw water prior to primary clarification to precipitate
phosphorus in the primary sedimentation basins. "Co-precipitation" involves adding chemicals directly
to the biomass in the biological system to form precipitates that can be removed with biological sludge.
"Post-precipitation" is when chemicals are added after secondary sedimentation and precipitants are
removed in a tertiary process such as sedimentation or filtration (Tchobanoglous et al. 2003). Chemicals
also can be added to waste or recycle streams to reduce the recycle of phosphorus released during
sludge handling. For example, at the Alexandria Sanitary Authority Advanced Wastewater Treatment
Plant, the alum in the return stream was reported to aid in phosphorus removal through the plant (EPA
Region 10 2007).
Nutrient Control Design Manual 9-4 August 2010
-------
.
>
k '
Primary
\
' ^
Ifc
Secon
1
^
Solids
Processing
1
— >
Figure 9-1. Possible application points for chemical addition (C).
Source: Adapted from Bott et al. (2007), with Permission from the Water Environment Research Foundation.
Table 9-2 summarizes the advantages and disadvantages of possible feed points for metal salts.
The optimal point of chemical addition is specific for a given plant and depends on the existing
treatment configuration, treatment goals, choice of chemicals, and solids handling issues. One of the
main advantages of multiple point addition is a savings in total chemical usage. The amount of metal salt
per unit of phosphorus removal increases as the final phosphorus concentration decreases. By using
multiple addition points, a large portion of the phosphorus removal can be done at a higher phosphorus
concentration and lower metal/ phosphorus removal ratio.
Lime cannot be added directly to biological treatment processes because it raises the pH above
10 for phosphorus removal and activated sludge processes prefers pH levels below 8.5. Lime can be
added to primary sedimentation tanks and removed with the primary sludge or it can be added as a
tertiary treatment process after biological treatment. When added to primary tanks, it will also result in
the removal of colloidal material through coagulation and settling, with a concomitant removal of TSS
up to 80 percent and COD up to 60 percent. In either case, the need for pH adjustment for downstream
processes should be considered and is typically accomplished by adding CO2 or a liquid acid such as
sulfuric acid, nitric acid, or hypochlorite (Tchobanoglous et al. 2003).
It is important to note that downstream biological treatment may be negatively affected if too
much phosphorus is removed by chemical addition in primary treatment, as phosphorus is an essential
nutrient for growth of microorganisms. For activated sludge, the minimum ratio of phosphorus to five-
day biochemical oxygen demand (BOD5) for a rapidly growing low solids retention time (SRT) system is
typically about 1:100 (WEF and ASCE 1998). Note, however, that the amount of phosphorus (and all
other nutrients) required is a function of net biomass yield and decreases as the operating SRT
increases.
Nutrient Control Design Manual
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Table 9-2. Advantages and Disadvantages of Metal Salt Application Points
Application Point
Advantages
Disadvantages
Primary Clarifier only
(pre-precipitation)
Removes additional BOD and
solids
Uses lower stoichiometric dose
Iron addition can reduce sulfide
odors
Reduces oxygen transfer
requirements in the biological
process, and reduces the amount
of excess biomass sludge
produced.
Control issue of leaving enough P for
biotreatment but low enough for effluent
Does not remove polyphosphates which
will be converted to orthophosphate in the
bioprocess
Competing reactions for hydroxides can
decrease dose efficiency
Removes alkalinity before nitrification
process, which can result in low pH levels
that inhibit nitrification
Removes BOD that can be used
downstream for denitrification. Can result
in larger anoxic tanks or an increased need
for an exogenous carbon source for
nitrogen removal.
Secondary Treatment
only, e.g., aeration
basin or before
secondary clarifier (co-
precipitation)
For effluent P less than 1.0 mg/L
good final control point for
chemical dosing
Polyphosphates converted so
most of P is available
May help improve TSS removal in
clarifiers
Help prevent fouling in MBR
systems
Removes alkalinity within the biological
nitrification process which can lower pH
and inhibit nitrification
MLSS increases with production of
chemical sludge, which increases the solids
loading to the final clarifiers. May need
larger activated sludge tanks or larger
clarifiers.
Tertiary Treatment only
(post-precipitation)
For effluent P less than 0.5 mg/L
good final control point for
chemical dosing
Polyphosphates already
converted so most of P is
available
Will help improve TSS removal
Can recycle precipitant to
headworks for added P removal
Filtration increases capital and operating
costs
Filtration increases operational complexity
and maintenance
Filter solids breakthrough can lead to
spikes in effluent P
P removal to low levels can inhibit or
prevent nitrogen removal by
denitrification filters
Requires separate sludge handling
Multiple
Can achieve lower effluent TP
concentration
Optimization of chemical dose to
lower requirements
Good control point at final dosing
Provides flexibility
• Additional costs for chemical feed and
control equipment in multiple locations.
• Additional operational complexity
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9.4 Determining the Chemical Dose
Chapter 3, Section 3.3.2 provides a detailed discussion of the reactions between phosphorus
and chemical precipitants and of dosing theory. For phosphorus removal using metal salts, the molar
ratio of the metal to the soluble phosphorus concentration (Med0se / PM) is the basis for the chemical
dose. As noted in Chapter 3, recent research by Smith et al. (2007) showed that for typical influent
phosphate concentrations.
• Doses above 1.5 to 2.0 Med0se/Pini ratios are sufficient to remove 80 to 98 percent of soluble
phosphorus.
• Reaching very low limits requires significantly higher ratios on the order of 6 to 7 Medose/Pmi-
See the example on the next page for an approach to calculating an initial ferric chloride dose using
these guidelines.
For phosphorus removal using lime, the dose is generally independent of the phosphorus
concentration because the lime first reacts with alkalinity to precipitate hydroylapatite (see Chapter 3
for more information). The recommended dose of lime is typically 1.4 to 1.6 times the total alkalinity
expressed as CaCO3 (Tchobanoglous et al. 2003).
Nutrient Control Design Manual 9-7 August 2010
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Example 9-1: Determining Initial Ferric Chloride Dose for Phosphorus Removal
Problem: A WWTP operator wants to use ferric chloride (FeCI3) to remove phosphate in the
secondary clarifier. The plant's flow rate is 1.0 MGD, and the measured phosphate concentration
entering the secondary clarifier is 4 mg/L as P.
Assumptions and Constants:
Ferric chloride solution strength = 40 percent
Density of ferric chloride solution = 1.4 kg/L
Molecular weight of Fe = 55.85
Molecular weight of FeCL3 = 162.2
Molecular weight of P = 30.97
Solution:
1. Determine the weight of iron (Fe) available per liter of liquid ferric chloride.
a. The weight of FeCI3/L per liter of ferric chloride solution is:
FeCI3/L = (0.40)(1.4 kg/L) = 0.56 kg/L
b. The weight of Fe per liter of ferric chloride solution is:
Fe/L = FeCI3/L x molecular weight of Fe / molecular weight of FeCI3
= (0.56 kg/L) x (55.85/162.2) = 0.193 kg/L
2. Determine the weight of Fe required per unit weight of P.
a. Target dosage = 2.0 mole Fe (Fedose) per 1.0 mole P (Pini) to remove up to 98 percent of
phosphate
b. Iron required = 2 x (Fedose/Pini)
= 2 x (55.85/30.97) = 3.61 kg Fe/kg P
3. Determine the amount of ferric chloride solution required per kg P.
Ferric chloride dose = 3.61 kg Fe x L ferric soln = 18.70 L ferric soln/kg P
1 kg P 0.193 kg Fe
4. Determine the amount of ferric chloride solution required per day
a. Convert 1.0 MGD to L/d = 1 x 10s gal/d x 3.785 L/gal = 3.785 x 10s L/d
b. Ferric chloride solution /d = (3.785 x 10s L/d) x (4 me P/L) x (18.65 L ferric soln/kg P)
(1 x 10s mg/kg)
= 283 L ferric chloride solution/d x 1 gal/3.785L
= 74.8 gal/d
Nutrient Control Design Manual 9-8 August 2010
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While the literature provides good general guidelines on chemical doses for phosphorus
removal, it does not account for site specific factors such as the influence of competing reactions and
the effects of pH and alkalinity. Thus, designers should conduct jar testing to determine the optimum
dose and mixing conditions for the specific wastewater to be treated. Jar testing is a fast, low-cost
procedure that is commonly used in the water and wastewater industry. It simulates rapid mixing and
flocculation on a small scale to predict the performance of large scale treatment operations. Figure 9-2
shows a common jar test apparatus. Multiple tests can be conducted at the same time to compare
alternative doses, pH ranges, or mixing speeds.
water
containers
mixing paddles
Figure 9-2. Schematic of common jar testing apparatus.
Source: Poland and Pagano, 1998, http://www.cee.vt.edu/ewr/environmental/teach/wtprimer/iartest/iartest.html
Although jar testing is common, careful planning is still required. Keys to successful jar testing
are summarized in the text box below.
Keys to Successful Jar Testing
• Determine variables to be tested (i.e., effect of temperature, pH,
mixing intensity, etc.)
• Vary only one parameter (e.g., alum dose) at a time. All of the
other variables such as mixing rate and contact time should remain
constant during a given test.
• Be careful to mix samples thoroughly prior to testing.
• Select reaction and flocculation times to mirror operational
conditions at the plant.
Source: adapted from WERF 2008
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9.5 Designing a Chemical Feed System
The objective of any feed system is to add the chemical to the wastewater at a controlled rate.
Designers should always consider the properties of the chemical being added (as listed in the Material
Safety Data Sheet) and state requirements when designing the chemical storage and feed system.
Safety precautions such as containment in case of spill and personnel safety equipment requirements
are extremely important.
Table 9-3 summarizes the types of chemical feeders typically used at municipal WWTPs. Section
9.5.1 provides guidance for design of liquid feed systems, followed by Section 9.5.2 with guidance on the
design of dry feed systems, including lime.
Table 9-3. Types of Chemical Feeders
Category
Dry Feeder:
Volumetric
Dry Feeder:
Gravimetric
Solution
Feeder:
Nonpositive
Displacement
Solution Feed:
Positive
displacement
Proportional
Pump
Feeder Type
Oscillating plate
1 Oscillating throat
j (universal)
Rotating disc
I Rotating cylinder
; (star)
Screw
] Ribbon
I Belt
j Continuous — belt
i and scale
: Loss in weight
: Decanter (lowering
pipe)
Orifice
Rotameter
(calibrated value)
I Loss-in-weight (tank
I w/control valve)
Rotating dipper
, Diaphragm
I Piston
Use Equipment Limitations
", General
Any material, granules, or
powder •
i Any material, any particle size I
Most materials including NaF, Use disk unloader
granules, or powder for arching
Any material, granules, or •
• powder I
Dry, free flowing material,
• powder, or granular •
' Dry, free flowing material, ]
• powder, granular, or lumps
j Dry, free flowing material up 1
j to 1.5-in size, powder, or 1
1 granular I
I Dry, free flowing granular I Use hopper or
• material, or floodable j agitator to
material j maintain
• Most materials, powder, ; constant density
, granular, or lumps •
: Most solutions or light :
slurries
Most solutions No slurries
Clear solutions No slurries
| Most solutions i No slurries
Most solutions or slurries
, Most solutions, Special unit .
" for 5% slurries3
Most solutions, light slurries |
Capacity
(mVh)1
0.001-3.1
0.002 - 9.0
oTooi-o^o9
0.7-180
0.65-27.0
0.005-1.7
0.002-0.015
0.009 - 270
0.002-0.18
0.002-7.2
0.0009-0.9
0.015-0.045
0.0005 - 0.015
0.0002 - 0.018
0.0002 - 0.018
0.009-2.7
0.0004-0.014
0.0001-15.3
", Feed rate
range2
: 40 to 1
• 40 to 1
2o"to"l
; 10 to 1 or
} 100 to 1
20tol
; 10 to i
I 10 to 1
I 100 to 1
; 100 to 1
: 100 to 1
lOtol
lOtol
| SOtol
100 to 1
! 100 to i
I 20 to 1
1. Volumetric feed capacities are given because chemical specific gravities must be known to specify mass feed capacity.
2. Ranges apply to purchased equipment. Overall feed ranges can be extended more.
3. Use special heads and valves for slurries.
Source: WEF and ASCE 2009, Table 16.13
Nutrient Control Design Manual
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9.5.1 Liquid feed systems
Liquid or solution feed systems are typically used in the following cases (WEF and ASCE 1998):
• If chemicals are applied at low rates.
• If chemicals are less stable.
• When the dust of the dry form of the chemical is hazardous or difficult to handle.
• When only liquid chemicals are available.
9.5.1.1 Storage
Alum is typically stored without dilution at the shipping concentration received at the plant.
Storage tanks located outside should be closed and vented, with provisions for heating to maintain
temperatures above -4°C (25°F) to prevent crystallization. Liquid alum storage vessels are constructed
of type 316 stainless steel, fiberglass-reinforced plastic (FRP), steel lined with rubber, or polyvinyl
chloride (PVC). Liquid alum can be stored indefinitely.
Ferric compounds are acidic and require special storage and handling procedures. Fiberglass-
reinforced plastic or polyethylene tanks can be used to store liquid ferric solutions (WEF and ASCE
2006).
Storage tanks should be sized to accommodate a 10-day to 2-week supply and should be
capable of handling 1-1/2 times the maximum quantity shipped. Some state regulations require that
storage times be related to potential delivery delays or specific periods at average chemical feed rates.
9.5.1.2 Feed Methods
Several alternatives are available for feeding liquid chemicals or chemical solutions. Descriptions
of common feed systems are provided below. Selection of the feed method is site specific and depends
on factors such as feed pressure, fluid properties, type of control, and treatment goals.
Manufacturers' recommendations should be followed regarding selection of pump materials for
the specific chemical of interest. Because iron and aluminum salt solutions are acidic and corrosive,
WEF and ASCE (2006, p. 297 and 303) recommend that pump heads be polyvinyl chloride (PVC) and that
piping, fittings, and valves be PVC or chlorinated PVC (CPVC). Pipe selection for polymer service should
be made after the type of polymer has been determined (plastic pipe or type 316 stainless steel is
normally used). Pipes that are above ground in temperate regions may need to be heat-traced.
It is important to note that it is generally not necessary to dilute liquid alum or ferric chloride
prior to feeding into the process. Addition of carrier water could raise the pH and result in formation of
metal hydroxides which can precipitate and cause plating in chemical feed lines. If dilution is needed,
carrier water should be added as close to the injection point as possible (WEF and ASCE 2006). For
polymers, dilution of the stock solution is often practiced to allow better dispersion of the polymer in
the wastewater.
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Rotary Dipper Feeders or Rotameters
Rotary dipper feeders are reliable feeders that are commonly used for gravity flow applications.
Feed rates can be varied based on a signal from a mainstream flow meter (flow proportional control).
Rotameters in conjunction with control valves may also be used for small applications where frequent
variation in chemical feed rate is not required. Rotameters should not be used with ferric chloride or
other iron solutions because the sight glass will become stained and opaque.
Non-positive Displacement Pump
A non-positive displacement pump, or a kinetic pump, delivers water in a steady stream. Energy
added to the fluid increases the flow velocity inside the pump relative to the velocity at discharge.
Examples of kinetic pumps include solids-handling centrifugal pumps, recessed-impeller pump,
screw/combination centrifugal pumps, and grinder pumps (WEF and ASCE 1998).
A centrifugal transfer pump is an example of a non-positive displacement pump. These pumps
should be directly connected but not close-coupled to prevent leakage into the motor
Positive Displacement Pump
When the discharge volumes of a pump are separated by a period of no discharge, it is referred
to as a positive displacement pump. A positive displacement pump pulls water into the pump chamber
with a vacuum created by the withdrawal of a piston, which displaces a set volume of water from the
chamber and forces it through the discharge valve and pipe (WEF 2008). Examples of positive
displacement pump types are the plunger pump, the reciprocating piston, peristaltic pump, pneumatic
ejector, rotary lobe pump, air-operated diaphragm pump, and progressing cavity pump (WEF and ASCE
1998). Metering pumps are either of the diaphragm or plunger type. Diaphragm pumps protected with
internal or external relief valves are preferred. A back pressure valve is recommended to provide
positive check valve operation.
Proportional Pumps
A proportional pump is a type of positive displacement pump commonly used for chemical feed.
It is a type of diaphragm pump that has a flow rate that can be precisely adjusted via changing the
length of the stroke or the stroke speed. The mechanism of a diaphragm pump is that the diaphragm is
a part made of flexible rubber-type material fastened within a vertical cylinder. The diaphragm is raised,
producing suction, and when it is dropped, it forces liquid out of a discharge valve (WEF 2008). It has
been reported that this type of pump is very susceptible to clogging and that a filter should be installed
to prevent any sediment from reaching the pump.
9.5.2 Dry Feed Systems
9.5.2.1 Storage
Bulk dry alum can be transferred with screw conveyors, bucket elevators, or pneumatic
conveyors. Bags and drums of alum should be stored in dry locations. Day hoppers receiving alum from
bags or drums should have a minimum storage capacity of eight hours at the maximum expected feed
rate. Hopper bottoms should have a minimum wall slope of 60 degrees to prevent arching. A typical
bulk storage tank for dry chemicals is shown in Figure 9-3.
Nutrient Control Design Manual 9-12 August 2010
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Dusl Fill Pipe
ColtectQc (Pneumatic)
Dust
Collector
Screen
wllTl
Breaker
/
Day Hopper tor
Dry Chemical
from Bags
or Drums
JJ
Alternative Supplies Depending on Storage
fe
Water *"
Supply
Dust and Vapor
Rernovor
1
Feeder
Dram
Rotameter
Control Solera*
Valve Valve
Scale or
'Sample
Chute
Baffle
i
Gravity to Aoolicaliofi ^
Levet
•L—JProtas
I
HoWino
{. ] "~
, Pressure
educing
Valve
Figure 9-3. Typical dry chemical feed system.
Source: USEPA (1987b),cited in Tchobanoglous et al. (2003)
9.5.2.2 Feed Methods
The following strategy for feeding applies to dry chemicals:
• Water is blended with the chemical in a mechanically-agitated dissolving tank or solution
tank.
• For bulk chemicals, a water meter in conjunction with a variable rate feeder achieves a
continuous stream of the solution at the proper strength.
• For bags or containers, the proper solution is made up manually on a batch basis.
Depending on the type of chemical, various feeding protocols are necessary:
• Alum, sodium aluminate or dry ferrous sulfate -
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/August 2010
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Recommended minimum solution strength is 6 percent or 0.06 kg/liter of water (0.5
Ib/gal).
Detention time in the dissolver should be 5 minutes at the maximum feed rate.
• Ferric sulfate-
Solutions are made up at a water-to-chemical weight ratio of 2:1 to 8:1, with a typical
ratio being 4:1 or 0.25 kg Fe2S04/Liter of water (2.1 Ib/gal).
Solutions under 1 percent strength are subject to hydrolysis and deposition of ferric
hydroxide.
The degree of automation in dry chemical dissolution systems will depend on the size of the
plant and daily chemical usage. For plants treating less than 1 million gallon per day (MGD), manual
preparation of the chemical solution on a batch basis may be sufficient. This is typically accomplished in
a day tank in which dry, bagged chemical is mechanically mixed with water to reach the desired
concentration. For larger facilities, the chemical solution is prepared automatically using a controller
that adjusts feed rate of dry chemical in proportion to potable water flow rate (as shown in Figure 9-3 in
Section 9.5.2.1) (Daniels 1973).
Volumetric feeders are the least expensive option and can be used where cost is a concern,
chemical delivery rates are low, and great accuracy is not required. Volumetric feeders generally
employ a screw feed mechanism.
Loss-in-weight gravimetric feeders provide a high degree of accuracy (up to 99 percent accurate)
and are recommended where close control of chemical dosages can result in substantial savings in
chemical costs. Belt gravimetric feeders are intermediate in cost between volumetric and loss-in-weight
gravimetric feeders, and can provide accurate and reliable service.
In general, closed construction is preferable for chemical feeders, because this exposes a
minimum of operating components to the corrosive vapors from the dissolving or solution tank.
Gravimetric feeders offer the following advantages over volumetric feeders:
• Calibration usually not required
• Greater accuracy and dependability
• Incorporation of totalizer to allow maintenance of accurate records and inventories
• Automatic proportioning
• Low maintenance; simple operation
9.5.2.3 Lime Slaking
Lime is available either as powdered hydrated lime or as pebble quicklime. Powdered hydrated
lime tends to be more convenient because it has already gone through the slaking process. Slaking is a
messy process that requires special considerations because of the inherent nature of lime:
• The reaction of lime and water generates a large amount of heat.
• Caustic properties create a hazard to operators and equipment.
Nutrient Control Design Manual 9-14 August 2010
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• Chemical handling creates lime dust.
• Lime has a tendency to precipitate out during slaking.
• Lime can cake when exposed to moisture and carbon dioxide in air.
• Lime can form a hard scale on equipment.
Due to the added cost of slaking, hydrated lime costs about 30 percent more than quicklime and
can produce extra dust during handling. However, the system can avoid the burden of slaking, and
hydrated lime tends to have a more stable storage life. For plants that use at least 2 to 3 tons of lime
per day, bulk quicklime may be the most affordable option. The best quality of quicklime has a high
percentage of purity (greater than 62 percent quicklime) and quick slaking ability. For plants where
lime use is not required daily, hydrated lime may be better in terms of convenience and storage (WEF
and ASCE 1998).
9.6 Designing for Rapid Mix and Flocculation
Rapid mix is a brief process where the chemical dose is vigorously blended with wastewater.
Chemicals should be dispersed uniformly to ensure formation of precipitates. Recent research has
confirmed that rapid mixing at the point of chemical dosage is critical for efficient phosphorus removal
using iron salts (Smith et al. 2007).
The rapid mixing process is followed by flocculation (slow mixing) to allow agglomeration of
solids to form larger floes with improved settling characteristics. It is important to minimize agitation of
the water during flocculation to prevent floe destruction. Existing plant components such as aerated
grit chambers, aerated distribution channels, or feed wells of clarifiers are commonly used for
flocculation, often after some modification.
This section provides information on mixer configurations available for rapid mix and
flocculation and guidance on setting design parameters.
9.6.1 Types of Mixers
Mixers are designed to achieve the goals of the type of mixing intended. Rapid mixing is a
shorter, faster process meant to maximize contact of coagulants and polymer with the wastewater,
while flocculation requires longer hydraulic retention times (HRTs) and more careful agitation to
increase floe size and settling. Typical mixer designs for both rapid mix and flocculation are described in
Table 9-4.
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Table 9-4. Commonly Used Equipment for Rapid Mixing
Type of
Mixing
Mixing
Equipment
General Description
Design Options
Rapid Mix
Static in-line mixer
Unique in that they have no moving
parts and provide mixing by using vanes
or plates to cause sudden changes in
flow direction. Most common type of in-
line mixer.
Mixing time varies with length of
mixer. Degree of mixing related to
pressure drop through the mixer.
Better mixing results in larger pressure
drops.
In-line mixer
Similar to static in-line but contain
rotating mixing element to enhance
process.
Power can be provided by an external
source or by the turbulence created by
the configuration of the mixer and
rotating element.
High-speed
induction mixer
An impeller is rotated at a sufficient
speed to cause a vacuum behind the
impeller. This vacuum pressure is then
used to draw chemicals to the impeller.
High speeds can result in
instantaneous dispersion of chemicals.
Pressurized water
jets
Chemical is added through a high
pressure water jet. The velocity of the
jet must be high enough to mix water in
all parts of the pipeline.
A small diameter reactor tube can
increase mixing effectiveness; power
provided by solution feed pump.
Rotating impeller
(Turbine or
propeller)
Constructed with vertical shaft driven by
electrical motor. Impellers mounted on
the shaft provide mixing. Flow can be
horizontal or vertical. Turbines are, in
general, higher velocity, and greater
turbulence leads to improved mixing.
High-speed mixers with small impellers
are better for rapid mix, while slow-
speed mixers with larger impellers are
used forflocculation.
Two basic impeller types are (1) radial-
flow impellers with flat or curved
blades mounted parallel to the axis of
the shaft, and (2) axial-flow impellers
at an angle < 90° with the drive shaft.
Axial-flow impellers can be variable
pitch-constant angle or constant pitch-
variable angle.
Small mixers revolve at 1750 rpm,
large mixers revolve at 400 to 800
rpm; more than one propeller may be
mounted on the same shaft.
Flocculation
Static
Wastewater is subjected to reversals in
flow
May include flow channels laid out
horizontally to cause energy-producing
friction; channel spacing may be
designed to decrease energy over time
to keep large floes from breaking apart
Paddle
Consists of a series of paddles spaced
apart and mounted to a shaft; slow
moving paddles promote gentle
flocculation. Less commonly used than
other methods because of maintenance
Paddles mounted on horizontal or
vertical shaft; agitation speed
important to promote floe formation
and maintaining floe size for
settleability.
Turbine and
Propeller
3 to 4 blades mounted to a vertical shaft.
Blades may be rectangular or hydrofoil
shaped; specific design meant to
reduce floe shearing; consider size
related to power and pumping; tip
speed and superficial velocity.
Ballasted
Uses continuously recycled media and a
variety of additives to improve settling
properties of suspended solids through
improved floe bridges.
Examples of proprietary designs
include Actiflo®, DensaDeg®, CoMag®
and Lamella® plate clarification
system.
Source: Tchobanoglous et al. (2003); USEPA (2003)
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August 2010
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Often, movement of wastewater through the plant is sufficient for flocculation, although care
should be taken not to disrupt floe formation through pumping or aeration. Another approach is to use
a flocculating clarifier, in which an expanded center well provides the desired detention time for slow
mixing. The contents of the flocculation well can be agitated by mechanical mixers or diffused air,
although the hydraulic regime in the center well may be such that mechanical or air mixing does not
provide additional benefit. Although not common, mechanical flocculation using a mixer or air agitation
can be used to improve performance of clarifiers. Proprietary technologies using ballasted flocculation
combine rapid mix, flocculation, and sedimentation in one unit. They can significantly reduce footprint
requirements and can be very effective at phosphorus removal. Chapter 6, section 6.3.1 provides a
description of several proprietary technologies.
9.6.2 Design Factors
The most important parameters for designing rapid mix and flocculation systems are the
velocity gradient (G), the hydraulic detention time (t), and the vessel geometry. For static mixers, the
degree of mixing is related to the headless through the mixer. For any kind of rotating impeller system
(for rapid mix or flocculation), the important design parameters are the rotational speed (n) and the
ratio of the impeller diameter to the equivalent tank diameter (D/Te). For paddle flocculators, the
power is related to the drag force on the paddles. Each of these design factors is discussed below. A
summary of key parameters is provided in Section 9.6.3. For additional guidance and example
calculations, see Chapter 5 of Tchobanoglous et al. (2003) and WEF and ASCE (1998).
9.6.2.1 Velocity Gradient
The velocity gradient is a measure of the shear intensity imparted to a fluid. It can be estimated
using the following equation (Tchobanoglous et al. 2003):
G = [P/Vu.]1/2 Eq. 9-1
Where:
G = Velocity gradient, second -1
P = Power requirement, W (kg m2 s"3)
u. = Absolute fluid viscosity, kg m"1 second"1
V = Basin volume, m3
For rapid mix, Smith et al. (2007) recommends G values greater than 300 seconds ~1, although
velocity gradients of up to 1,000 seconds -1 have been recommended (Barth and Stensel 1981). Velocity
gradients for flocculation processes generally are 30-60 seconds ~1, depending on the chemicals added
and point of addition (Tchobanoglous et al. 2003). Lower velocity gradients may yield floe particles with
too much trapped water, whereas higher velocity gradients may cause excessive floe shear and floe
deterioration.
In some cases, chemicals are added directly to the activated sludge process with no additional
mixing other than that already designed for aeration and mixing of solids and mixed liquor in the basin.
Although this practice can be effective, it represents less than ideal conditions for flocculation, as
velocity gradients in aeration basins usually result in floe shear. In one study, air flowrates in the
downstream end of the aeration basin were reduced to achieve a velocity gradient of 60 seconds ~1,
Nutrient Control Design Manual 9-17 August 2010
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which was found to be optimum for flocculation using ferric chloride (Singhal 1980). Addition of anionic
polymer prior to clarification assists in the agglomeration of sheared floe.
9.6.2.2 Power Requirements
If a target velocity gradient has been selected, designers can derive the theoretical
requirement using equation 9-1, solving for P as follows:
power
P = G2 u.V
Eq. 9-2
For rotating impeller systems, the power is related to the revolutions per second and diameter
of the impeller by the following equation:
Eq. 9-3
Where:
P = Np p n3 D5
P = Power input, W
Np = Power number for the impeller, unitless
p = Mass density of the fluid, kg/m3
n = Impeller revolutions per second, s"1
D = Diameter of the impeller, m
Pumping capacity of the mixer can be estimated using similar inputs:
Qi = NQnD"
Eq. 9-4
Where:
Q, = pump discharge, m3/sec
NQ =flow number of the impeller, unitless
Power numbers and flow numbers for different types of impellers along with qualitative information on
pumping capacity are provided in Table 9-5. Designers should consult manufacturers for power and flow
numbers for a specific piece of mixing equipment.
Table 9-5. Values of NP and NQ for Various Types of Impellers
Type of Impeller
Vertical flat-blade turbine (VBT)
Disk turbine
Pitched-blade turbine (45° PBT)
Pitched-blade turbine (32° PBT)
Low-shear hydrofoil (LS, 3-blade)
Low-shear hydrofoil (LS, 4-blade)
Propeller
NP
3.5-4
1.6
1.1
0.30
0.60
NQ
0.84-0.086
0.084-0.086
0.084-0.086
0.50
0.55
Pumping Capacity
Low
Low
Moderate
High
High
High
Source: Tchobanoglous et al. (2003), Table 5-12
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Equation 9-3 applies if the flow is turbulent (i.e., the Reynolds number, NR, is greater than
10,000), which is generally the case for rapid mix designs. When mechanical paddle mixers are used for
flocculation, power is related to the drag force on the paddles as follows (Tchobanoglous et al. 2003):
P = (CDApv2p) / 2 Eq. 9-5
Where:
P = Power requirement, W
CD = Coefficient of drag of paddle moving perpendicular to fluid, unitless
A = Cross sectional area of the paddles, m2
p = Mass density of the fluid, kg/m3
vp = Relative velocity of the paddles with respect to the fluid, m/s, usually assumed to be 0.6 to
0.75 times the paddle-tip speed.
This equation can be rearranged to determine the required paddle area for the required power, as
derived using equation 9-2.
For static mixers, the key parameter for design is not power required since there are no moving
parts, but headless through the mixer. The headless can be estimated as follows:
h = KSMv2 Eq. 9-6
Where:
h = Headless across the static mixer, m
KSM = Overall coefficient for the static mixer, sec2/m, as provided by the manufacturer
v = Approach velocity, m/sec
9.6.2.3 Hydraulic Retention Time
The hydraulic retention (or detention) time is the amount of time a unit of the process flow
stays within a tank for a given flowrate. It is calculated as follows:
t = V/Q Eq. 9-7
Where:
t = Hydraulic retention time, seconds
V = Reactor volume, ft3 or m3
Q = Flow rate, ft3/sec or m3/sec
Depending on the mixing device and chemical coagulant chosen, a chemical addition process
will require different rates of rapid mixing to ensure adequate hydraulic retention time, but still limit
contact in the mixer. For example, static in-line mixers, in-line mixers, and high-speed induction mixers
are used for alum, ferric chloride, and cationic polymer when instantaneous mixing is required. A
hydraulic retention time of only 1 second or less is required for these types of mixers because the
chemical precipitants tend to have a fast reaction rate. However, applications using lime may require
between 10 and 30 seconds of hydraulic retention time since lime reacts more slowly with wastewater
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(Tchobanoglous et al. 2003). Ten State Standards (2007) recommend a detention time of at least 30
seconds for basins equipped with mechanical mixing devices.
The process of the flocculation of small particles of solid phosphorus is not completely
understood. These particles are affected by pH and the surface chemistry of the metal-to-phosphorus
complexes (WERF 2008). Smith et al. (2008) reported that floe aging may contribute to a decrease in the
phosphorus removal efficiency of wastewater. Spent floes can be recycled to promote flocculation in
the wastewater stream, but based on this research, the effect of floe age on decreasing floe surface area
and the related ability to bind with phosphorus should be considered. Despite the somewhat lesser
understood facts about flocculation, it is commonly known that polymer can improve flocculation. The
polymer works by capturing small colloidal particles and helping them to join via the long chain polymer
molecules to form larger particles that will settle out more readily.
9.6.2.4 Vessel Geometry
Vessels for rapid mixing should be designed to avoid the formation of vortexes. If the vessel is
small, the impellers should be mounted off-center or at an angle. For larger vessels, designers
commonly avoid vortexing by installing four or more vertical baffles (also called stator baffles) extending
approximately l/10th to l/12th of the diameter out from the wall (Tchobanoglous et al. 2003). Vortexing
can also be minimized by choosing a flat bladed impeller instead of a fan or propeller impeller and
introducing chemicals at the blade level.
The ratio of the impeller diameter to the effective tank diameter (D/Te) is an important
parameter for design and depends on whether the flow moves horizontally or vertically through the
mixing vessel. Tchobanoglous et al. (2003) recommend D/Te values of 0.25 to 0.4 for horizontal flow
vessels and 0.40 to 0.60 for vertical flow vessels. Note that Te for rectangular basins can be estimated as
follows:
Te = 1.13(LXW)1/2 Eq. 9-8
Where:
Te = Effective diameter for a rectangular tank, m
L = Basin length, m
W = Basin width, m
9.6.3 Summary of Typical Design Parameters
Table 9-6 summarizes typical design parameters for rapid mix. These values should be used as
general guidelines; final design values for velocity gradient and hydraulic retention time should be
established during jar or pilot testing and based on manufacturers' recommendations. Design
parameters for ballasted systems are provided in the EPA Wastewater Technology Fact Sheet, Ballasted
Flocculation (USEPA 2003), available online at
http://www.epa.gov/owm/mtb/ballasted flocculation.pdf.
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Table 9-6. Typical Design Parameters for Turbine and Propeller Mixers
Parameter
Symbol
Unit
Value
Horizontal-flow mixing
Velocity gradient
Rotational speed
Ratio impeller diameter to
equivalent tank diameter
G
N
o/re
l/s
rpm
unitless
500-2500
40-125
0.25-0.40
Vertical-flow mixing
Velocity gradient
Rotational speed
Ratio impeller diameter to
equivalent tank diameter
G
N
o/re
l/s
rpm
unitless
500-2500
25-45
0.40-0.60
Source: Tchobanoglous et al. (2003), Table 5-11
9.7 Solids Separation Processes
Although chemical addition remains a reliable method for phosphorus removal, there have been
few dramatic changes in both the types of chemicals available and their applications. However, to fill
the need for lower and lower effluent phosphorus requirements, the field of solids separation has
expanded to provide more effective options for removing insoluble phosphorus.
The options for solids separation depend on a multiple barrier approach—stand alone processes
may be used for solids separation, but depending on treatment goals, two or more systems together
may be able to achieve lower and more reliable phosphorus concentrations. In general, the solids
separation process ranges from the conventional to the advanced as follows (WERF 2008):
• Conventional sedimentation uses clarifiers for settling of floe, and, much of the time, follows
that with the secondary barrier of a filtration process to remove additional solids.
• High rate flocculation and sedimentation includes a more advanced clarifier to process larger
volumes of wastewater more quickly (e.g. ballasted flocculation/sedimentation, lamella® tube
settlers). Commonly followed by filtration.
• The direct filtration method uses a single barrier approach with no sedimentation.
• Two-stage filtration uses multiple filters.
• Microfiltration or ultrafiltration theoretically removes all insoluble phosphorus since the filter
size is under the 0.45 u.m usually considered to be the limit of insolubility; in practice,
membrane defects can lessen filter effectiveness.
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• Ballasted separation and magnetic polishing, such as the patented CoMag® process discussed
later in this section.
9.7.1 Primary and Secondary Clarification
Solids separation processes can be applied at various locations in the process train, typically
directly following chemical addition. When chemicals are added prior to primary or secondary
clarification, clarifier design becomes a key factor in removal of the precipitated phosphorus. Clarifiers
used in chemical precipitation systems differ little from those employed in conventional biological
treatment, although use of flocculation zones is recommended to allow flocculation to occur after
addition of coagulants. Provision of distinct flocculation zones is recommended for either primary or
secondary clarifiers, depending on the point of chemical addition. Distinct flocculation zones are
particularly important for primary clarifiers, because there may be little opportunity for flocculation to
occur in existing processes. For secondary chemical addition, flocculation can occur in aeration basins or
channels preceding clarification, but the use of flocculation zones in secondary clarifiers is a
recommended practice as it allows flexibility in the point of chemical addition and provides a zone in
which direct control can be exercised over velocity gradients to achieve optimum flocculation.
The WEF Manual of Practice FD-8, Clarifier Design (WEF 2005) provides detailed design guidance
for achieving solids removals in clarifiers including recommended design standards, software modeling,
details on processes and equipment, and performance monitoring and control. Bott et al. (2007) report
that solids removal through conventional clarifiers can reliably remove TP to effluent levels between 0.5
and 1.0 mg/L, and lower effluent levels can frequently be obtained.
9.7.2 Tertiary Processes
Tertiary processes (post-secondary treatment) use clarification, some form of filtration
technology, or both and can be used to consistently remove phosphorus to very low (< 0.1 mg/L)
concentrations. Tertiary clarifiers can be conventional, one or two sludge lime (second sludge includes
recarbonation and settling), solids contact, high rate, and ballasted-high rate. They are sized on the
same basis as primary and secondary units. High rate and ballasted high rate clarification is typically
combined with flocculation and rapid mixing in proprietary systems. These proprietary systems have
been described previously in Section 9.6. Chapter 6 summarizes tertiary filtration technologies with
design guidelines presented in Chapter 11 of this manual.
9.8 Operational Factors
9.8.1 Dose Control
The success of phosphorus removal by chemical addition depends on proper instrumentation
and control. Dosage control typically takes the form of manual operation (for small systems),
adjustments based on automatic flow measurements, or the more advanced on-line analyzers with
automated dosage control. Flow-paced dosing (feed-forward control) is common. More advanced
control systems may use online phosphate analyzers or oxidation reduction potential (ORP) for control.
See Chapter 13 of this design manual for additional information on control strategies.
Nutrient Control Design Manual 9-22 August 2010
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9.8.2 Make-up Water
Chemical properties of any water used for making solutions should be considered—tap water
high in dissolved solids could cause sludge to form when mixed with coagulants (WEF and ASCE 1998)
and could lead to clogging of chemical feed lines.
9.8.3 Sludge Production and Handling
Chemical precipitation methods always produce additional solids due to generation of metal- or
calcium- phosphate precipitates and additional suspended solids. Lime traditionally produces a higher
sludge volume compared to metal salts because of its reaction with natural alkalinity. An advantage of
lime sludge is that some stabilization can occur due to the high pH levels required. One disadvantage is
that lime can cause scaling in mechanical thickening and dewatering systems. Although it tends to
produce less sludge than do ferric salts, alum sludge can be more difficult to concentrate and dewater
compared to the ferric sludge.
Stoichiometric equations for estimating sludge production are provided in Section 3.5 of this
design manual, but a good rule of thumb is about 10 g of chemical sludge per g P removed by chemical
treatment. Example calculations are provided in the text box on the next page.
Nutrient Control Design Manual 9-23 August 2010
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Example 9-2: Determining Increase in Sludge Production from Addition of Ferric Chloride
in Example 9-1
Problem: A WWTP operator wants to use ferric chloride (FeCI3) to remove phosphate in the
secondary clarifier. The plant's flow rate is 1.0 MGD, and the measured phosphate concentration
entering the secondary clarifier is 4 mg/L as P. From Example 9-1 and confirmed through jar testing,
74.8 gallons of ferric chloride solution will be added per day to remove up to 98 percent of the
phosphate. The operator wants to estimate the mass of additional solids that will need to be
removed from the secondary clarifier as a result of the ferric chloride addition.
Constants:
Ferric chloride solution strength = 40 percent
Density of ferric chloride solution = 1.4 kg/L
Molecular weight of Fe = 55.85; Molecular weight of FeCL3 = 162.2; Molecular weight of P = 30.97
Molecular weight of O = 16; Molecular weight of H = 1
Solution:
1. Assume that the phosphorus removed will be contained in the following precipitate with iron as
Fei.6(H2PO4)(OH)3.8 and the remainder of the iron added will be as Fe(OH)3. See Chapter 3, Section
3.5 for additional discussion.
2. Calculate the millimoles per liter (mM/L) of iron added per day that precipitates as mM/L of
Fei.6(H2PO4)(OH)3.8 and as mM/L of Fe(OH)3.precipitate.
a. P removed as mM/L = (0.98)(4 mg P /L)/(30.97 mg P/mM P) = 0.127 mM P removed/I.
b. The amount of iron added was 2.0 mM Fe/mM P in feed. Fe added as mM/L = (2.0 mM
Fe/mM P)(4 mg P/L/(30.97 mg P/mM P) = 0.258 mM Fe/L
c. Precipitate as mM Fe/L in Fei.6(H2PO4)(OH)3.8 = 1.6 mM Fe/mM P removed = (1.6)0.127 mM P
removed/L = 0.203 mM Fe/L
d. Precipitate as Fe in Fe(OH)3 = mM/L of Fe added minus mM/L of Fe in Fei.6(H2PO4)(OH)3.8. =
0.258 mM Fe added/L - 0.203 mM Fe/L = 0.055 mM Fe in Fe(OH)3/L
3. Calculate sludge production as sum of Fei.6(H2PO4)(OH)3.8 and Fe(OH)3.
a. MW of Fe1.6(H2PO4)(OH)3.8 = 1.6(55.85) + 2 + 30.97 + (4)(16) + 3.8(16) + 3.8(1) = 250.9; MW of
Fe(OH)3 = 55.85 + (3)(16) + (3)(1) = 106.85
b. Sludge as Fe1.6(H2PO4)(OH)3.8 = 0.203 mM Fe/L( 1.0 mM Fe1.6(H2PO4)(OH)3.8/1.6 mM Fe)(250.9
mg/mM) = 31.83 mg/L
c. Sludge as Fe(OH)3 = 0.056 mM Fe/L(1.0 mM Fe(OH)3/mM Fe)(106.85) = 5.98 mg/L
d. Total sludge production from FeCI3 addition = 31.83 + 5.98 = 37.8 mg/L= (37.8 mg/L)(8.34
lb/MG-mg/L)(1.0 MGD) = 315.4 Ib/d (143.1 Kg/d)
The inorganic sludge production of 315.4 Ib/d (143.1 kg/day) is in addition to what would be
removed without the addition of ferric chloride. This is an initial design estimate and may vary
depending on wastewater conditions and the point of addition.
Nutrient Control Design Manual 9-24 August 2010
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9.8.4 pH Adjustment
Depending on the wastewater composition and pH, pH adjustment may be necessary to achieve
efficient phosphorus removal with chemical addition. Other uses for pH adjustment include the
neutralization of the wastewater flow either following chemical addition and prior to biological
treatment, or prior to effluent discharge. The important components of a pH adjustment process
include the wastewater flow rate, hydraulic retention time, and the precision of the chemical metering
system. Laboratory titration experiments are generally necessary to determine the correct dose of acid
or base (WEF and ASCE 1998). Target pH values for optimum results vary with the chemical being used,
and typical values are given in the literature. However, jar tests should be used to determine and
confirm the specific optimum value for the chemical and wastewater being reacted.
Raising the pH of an acidic wastewater stream is generally accomplished through the use of
lime, such as high calcium lime, quicklime, hydrated or slaked lime, or dolomitic lime. Calcium and
magnesium oxides or hydroxides may also be used in a mixture. Lime products tend to be insoluble and
require longer retention times than other chemicals—calcium and magnesium oxides in particular
require long HRTs, but are favored due to their relative low cost. As with other applications of lime, lime
used for pH adjustment creates a significant amount of sludge. Caustic soda and soda ash are much
more soluble, create less sludge, and react much more quickly with acid in wastewater. The higher cost
of caustic soda and soda ash compared to lime compounds should be factored into any decision-making
process. Dry soda ash can be fed by a continuous feeder using volumetric, gravimetric, or loss-in-weight
gravimetric mechanical feeders, while a solution feed can be pumped (WEF and ASCE 1998).
Lowering a high pH to a more neutral level for effluent discharge may take place via an acid feed
system that adds small quantities of sulfuric acid or hydrochloric acid to the wastewater. For larger
systems, the acid solution may be applied at the point of discharge. Nitric acid is readily available but
generally not preferred due to the addition of nitrogen species (WEF and ASCE 1998).
9.8.5 Effect on Biosolids Application
The use of metal salts can lead to large increases in inorganic solids in the sludge, which can
result in increased inorganic salts (salinity) in sludge and in the effluent. Salinity can create problems
when biosolids are land applied or when the effluent is returned to existing water supply reservoirs.
Biological phosphorus removal was developed in South Africa due to the high rate of indirect recycling
of wastewater effluent, which led to excessive total dissolved solids (TDS) during dry periods. High total
salts can reduce germination rate for crops and negatively affect the soil structure.
9.9 References
10 States Standards. 2007. Recommended Standards for Water Works, 2007 Edition: Policies for the
Review and Approval of Plans and Specifications for Public Water Supplies. Water Supply Committee
of the Great Lakes - Upper Mississippi River Board of State and Provincial Public Health and
Environmental Managers. Available online: http://10statesstandards.com/waterstandards.html
American Public Health Association (APHA), AWWA, and Water Environment Federation (WEF). 1998.
Standard Methods for the Examination of Water and Wastewater. 20th Edition. 220 pp. Washington,
D.C.: APHA, AWWA, and WEF.
Nutrient Control Design Manual 9-25 August 2010
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ASCE. 1977. Wastewater Treatment Plant Design. ASCE Manual No. 36. WPCF Manual of Practice No.
8, American Society of Civil Engineers, New York, NY and Water Pollution Control Federation,
Washington, DC, 1977.
Barth, E.F. and H.D. Stensel. 1981. International Nutrient Control Technology for Municipal Effluents.
JWPCF53(12):1691-1701.
Black, S.A. 1980. Experience with Phosphorus Removal at Existing Ontario Municipal Wastewater
Treatment Plants. In: Phosphorus Management Strategies for Lakes, eds: RC. Loehr, C.S. Martin, W.
Rast. Ann Arbor Science Publishers, Inc. Ann Arbor, Ml.
Bott, C.B., S. N. Murthy, T. T. Spano, and C.W. Randall. 2007. WERF Workshop on Nutrient Removal: How
Low Can We Go and What is Stopping Us from Going Lower? Alexandria, VA: WERF. ChemExpo. 2000.
Online Chemical Exposition, Archives. May 2000.
City of East Moline. 2007. Minutes of the Meeting of the Mayor and the Committee-of-the-Whole of
the City of East Moline, Illinois. November 19. Available online:
http://www.eastmoline.com/Minutes/COW/ll-19-
Q7.cfm?MonthChange=None&eventid=78&LastDate=08/31/2008
City of Santa Cruz. 2006. Ferric Chloride for the Wastewater Treatment Facility - Amendment #1 to
Sole Source Purchase Agreement. City Council Agenda Report. City of Santa Cruz, CA. February 9.
Gulp, R.L.., Wesner, G.M. and G.L. Gulp. 1978. Handbook of Advanced Wastewater Treatment. Second
Edition. Van Nostrand Reinhold Company, New York.
Daniels, S.L. 1973. Instrumentation and Automatic Control of Phosphorus Removal Processes. In:
Phosphorus Removal Design Seminar, Conference Proceedings No. 1. Toronto, Canada. May 28-29.
EPA Region 10. 2007. Advanced Wastewater Treatment to Achieve Low Concentration of Phosphorus.
EPA910-R-07-002.
Infilco Degremont, 2000. Design information on the DensaDeg system.
Gulbrandsen. 2008. Polyaluminum Chloride Solution Product Description. Clinton, NJ: Gulbrandsen
Technologies, Inc. Available online: http://www.gulbrandsen.com/pdf/PolyalumCI_4.pdf
Heim, N.E and B.E Burris. 1979. Chemical Aids Manual for Wastewater Treatment Facilities. NTIS No.
PB-116816, U.S. Environmental Protection Agency, Office of Water Program Operations, Washington,
DC. EPA-430/9-79 018.
Hortskotte, G.A., D.G. Niles, D.S. Parker, and D. H. Caldwell. 1974. Full-scale testing of a water
reclamation system. Journal of the Water Pollution Control Federation. 46(1): 181-197.
Miller, M.M. 2008. 2007 Minerals Yearbook: Lime [Advance Release]. U.S. Geological Survey. U.S.
Department of the Interior. Available online:
http://minerals.usgs.gov/minerals/pubs/commoditv/lime/mybl-2007-lime.pdf
Plank, D. 2008. "Chemical Prices Skyrocket, No End in Sight." E-Mainstream. AWWA. Posted 15 July.
Available online: http://www.awwa.org/publications/MainStreamArticle.cfm?itemnumber=39577
Nutrient Control Design Manual 9-26 August 2010
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Poland, J. and T. Pagano. 1998. Jar Testing. Water Treatment Primer, CE 4124: Environmental
Information Management. Civil Engineering Dept. Virginia Tech. Available online:
http://www.cee.vt.edu/ewr/environmental/teach/wtprimer/jartest/jartest.html
Reardon, R.D. 2005. Tertiary Clarifier Design Concepts and Considerations. Presented at WEFTEC 2005.
Saginaw City Council. 2006. Memo: Water Treatment Plant Liquid Ferric Chloride Purchase. Bids
received for chemical supply from July 1, 2006 -June 30, 2008. City of Saginaw, Ml. Available online:
http://www.saginaw-mi.com.
Singhal, A.K. 1980. Phosphorus and Nitrogen Removal at Cadillac, Michigan. JWPCF. 52(11):2762-2770.
Smith, S., A. Szabo, I. Takacs, S. Murthy, I. Licsko, and G. Daigger. 2007. The Significance of Chemical
Phosphorus Removal Theory for Engineering Practice. In Nutrient Removal 2007. WEF.
Szabo, A., I. Takacs, S. Murthy, G.T. Daigger, I. Licsko, and S. Smith. 2008. Significance of Design and
Operational Variables in Chemical Phosphorus Removal. Water Environment Research. 80(5):407-
416. Alexandria, VA: WEF.
Smith, S., I. Takacs, S. Murthy, G.T. Daigger, and A. Szabo. Phosphate Complexation Model and Its
Implications for Chemical Phosphorus Removal. 2008. Water Environment Research. 80(5): 428-438.
Alexandria, VA: WEF.
Tchobanoglous, G., F. L. Burton, and H.D. Stensel. 2003. Wastewater Engineering: Treatment and
Reuse. New York, NY: McGraw-Hill.
US Filter Kruger, 2000. Design information on the Actiflo® process for wastewater.
USEPA. 1975. Process Design Manual for Suspended Solids Removal. EPA-625/l-75-003a, NTIS No. PB-
259147, U.S. Environmental Protection Agency, Center for Environmental Research Information,
Cincinnati, OH, 1975.
USEPA. 1976. Process Design Manual for Phosphorus Removal. Great Lakes National Program Office.
GLNPO Library. EPA625/l-76-001a. April 1976.
USEPA. 1987a. Design Manual: Phosphorus Removal. Center for Environmental Research Information.
Cincinnati, OH. EPA/625/1-87/001.
USEPA 1987b. Retrofitting POTWs for Phosphorus Removal in the Chesapeake Bay Drainage Basin.
Cincinnati, OH. EPA/625/6-87/017
USEPA. 1999. Enhanced Coagulation and Enhanced Precipitative Softening Guidance Manual. Office of
Water. EPA 815-R-99-012.
USEPA. 2000. Wastewater Technology Fact Sheet: Chemical Precipitation. Office of Water. EPA 832-F-
00-018.
USEPA. 2003. Wastewater Technology Fact Sheet: Ballasted Flocculation. Office of Waste
Management. Municipal Technology Branch. EPA 832-F-03-010.
Nutrient Control Design Manual 9-27 August 2010
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USEPA. 2008. Emerging Technologies for Wastewater Treatment and In-Plant Wet Weather
Management. EPA 832-R-06-006. Available online:
http://www.epa.gov/OW-OWM.html/mtb/emerging_technologies.pdf
WEF. 1983. Nutrient Control. WEF Manual of Practice No. FD-7. Water Environment Federation.
Washington, DC.
WEF. 2005. Clarifier Design, 2nd Edition. WEF Manual of Practice No. FD-8, Water Environment
Federation, Washington, DC.
WEF. 2008. Operations Glossary. [Accessed 1 October 2008]. Available online:
http://www.wef.org/ConferencesTraining/SkillsBuilder/Operations_Glossary.htm
WEF and ASCE. 1998. Design of Municipal Wastewater Treatment Plants - MOP 8, 4th Ed. Water
Environment Federation and American Society of Civil Engineers. Alexandria, VA: WEF.
WEF and ASCE. 2006. Biological Nutrient Removal (BNR) Operation in Wastewater Treatment Plants -
MOP 29. Water Environment Federation and the American Society of Civil Engineers. Alexandria, VA:
WEFPress.
WEF and ASCE. 2009. Design of Municipal Wastewater Treatment Plants - WEF Manual of Practice 8
and ASCE Manuals and Reports on Engineering Practice No. 76, 5th Ed. Water Environment
Federation, Alexandria, VA, and American Society of Civil Engineers Environment & Water Resources
Institute, Reston, Va.
WERF. 2008. Nutrient Compendium. Chapter 1, Tertiary Phosphorus Removal. Available online:
http://www.werf nutrientchallenge.org/chapterl.asp?area=chl
Wilkes, W. 1973. Phosphorus Removal by Chemical Addition Using Primary Treatment. In: Phosphorus
Removal Design Seminar, Conference Proceedings No.l, Toronto, Canada, May 28-29.
YubaCity. 2007. Agenda for Staff Report on Chemicals (FB 07-07). City of Yuba City, CA. Available
online: http://www.vubacitv.net/documents/Agendas/2007/041707Council/041707Agendaltem9.pdf
Nutrient Control Design Manual 9-28 August 2010
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10. Design Approach for Biological Nutrient Removal
I
Chapter 10 covers:
I
10.1 Introduction
10.2 Preliminary Design Approach
10.3 Overview of Recommended Approach for Plant Modeling
10.4 Establishing Objectives and Requirements
10.5 Selecting a Process Simulator
10.6 Data Collection
I
I
i
10.7 Characterization of Organic Material
I 10.8 Characterization of Nutrient Fractions
10.9 Kinetic and Stoichiometric Parameters
10.10 Calibration
10.11 Validation
10.12 Simulation of Design Alternatives for Nutrient Removal
1 %
10.13 Additional Procedures for Design |
10.14 Design Checks for Nitrogen and Phosphorus Removal |
10.15 References |
10.1 Introduction
By this stage of the process, design engineers should have completed the following steps:
• Reviewed the mechanisms for biological nutrient removal at wastewater treatment plants
(WWTPs) (see Chapters 4 and 5).
• Characterized the wastewater and established design objectives (see Chapter 7).
• Reviewed flow sheets and other information on available treatment methods for biological
nutrient removal (see Chapter 6).
• Selected candidate processes for the plant upgrade (see Chapter 8).
The designer is now ready to select and design the final upgrade option (or start process selection for
Greenfield plants) and the operating conditions under which the plant can achieve desired objectives for
nutrient reduction.
Although process design for WWTP upgrades for biological nutrient removal can still be done
using traditional engineering methods (i.e., hand calculations or spreadsheets using theoretical
equations), process simulators1 have become increasingly powerful, easy to use, and widely accepted in
1 For the purposes of this design manual, "mathematical models" or more simply "models" are sets of equations that describe a
physical system. Examples include the activated sludge models (ASM), biofilm models, and computational fluid dynamic (CFD)
models for secondary clarifiers. "Process simulators" or "simulators" are computer programs that allow the user to specify all
the processes of a WWTP, select models for each, and solve the models to predict plant performance. Examples are BioWin
and GPS-X (See Table 10-1 for a full list)
Nutrient Control Design Manual 10-1 August 2010
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recent years. Simulators, which solve mathematical models and typically provide results in a user-
friendly format, allow designers to study kinetic- as well as time-based solutions while determining the
total mass balances of many constituents. WEF and ASCE (2010) note that "The complexity of biological
activity and reactor behavior and the number of variables important to nutrient removal require the use
of computer models for detailed solutions."
Simulators have many additional advantages over traditional design. Designers can use
simulators to quickly test many different treatment configurations and design parameters in a fraction
of the time and cost it would take to perform traditional design calculations and laboratory or pilot
tests. The additional cost of development (e.g., cost of setting up the simulator, special sampling to
define inputs) is usually small compared to the monetary benefits of more accurate predictions and
thus, lower safety factors in design. For these reasons, simulation has become the preferred method of
process design for biological nutrient removal at WWTPs.
Benefits of Using Models for Process Design
The traditional approach for designing biological nutrient removal systems (e.g., for nitrification) was
to determine solids residence time (SRT) based on kinetic equations, then to apply a safety factor
based on engineering judgment. The new method currently being used by design engineers is to
determine site specific nitrification kinetics based on influent quality and perform dynamic modeling
of the system to capture variability. Lower safety factors can be used in design with more confidence
that the plant can achieve the desired treatment objectives under a variety of operating conditions.
Although models are the preferred method for design of biological nutrient removal, there may
be instances where preliminary hand or spreadsheet calculations can be useful alone or as a check on
modeled output. Most engineering firms have used the mathematical equations found in Chapters 4
and 5 of this manual in a spreadsheet to estimate the basins sizes. Such spreadsheets can be developed
to estimate the following:
• Size of the anaerobic zone.
• Solids Residence Time (SRT) requirements for nitrification at minimum expected mixed liquor
temperature.
• Minimum requirements for denitrification zones.
• Mixed liquor recycle rates.
• Oxygen requirements for nitrification and carbon removal.
• Production of surplus, i.e. waste activated sludge (WAS).
• Clarifier size requirements.
• Return activated sludge (RAS) pumping ranges.
• Effect of approximate peaking factors.
Nutrient Control Design Manual 10-2 August 2010
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Section 10.2 provides a methodology for a preliminary design approach using equations and
graphs. The remainder of this chapter (Sections 10.3 through 10.14) presents the recommended
modeling approach, providing practical guidance on how to use commercially available process
simulators to design nutrient removal processes. This Chapter contains guidance on selecting a
simulator, establishing model objectives and requirements, collecting and verifying data, calibrating and
validating the model, and simulating alternatives. Examples are provided throughout.
This chapter assumes some familiarity with
process simulators; readers using models for the first p|gnt operatjon js Cfftfco/ to mode|jng nutrjent
time should carefully review technical remova| pmcesses Mjstakes jn
documentation and work with vendor characterization can have severe impacts on
representatives to get the simulator up and running desjgn ^ g gQod portjon Qf thjs chapter jg
properly. Readers are also encouraged to review devoted tQ provjdjng.
industry standards and technical documents for _ Gujdance Qn gatherjng jnput data
additional recommendations on modeling protocols,
wastewater characterization methods, and detailed _ |nformatjon on why these data are
calibration procedures. A particularly good important in the design of nutrient removal
reference is the 2003 Water Environment Research
Foundation (WERF) report, Methods for Wastewater
Accurate characterization of influent flow and
Discussion of how the data are related
Characterization in Activated Sludge Modeling (Melcer et al. 2003). In addition, modelers should check
for the pending publication of the International Water Association (IWA) Scientific Technical Report,
Good Modeling Practice (GMP) — Guidelines for Use of Activated Sludge Models (IWA 2009). Historical
background and information on underlying mathematical models is provided in Appendix C of this
manual.
10.2 Preliminary Design Approach
For preliminary design, a simple step-by-step approach (provided below) can be used to
estimate basin size for biological nutrient removal (nitrogen and phosphorus). Note that this
methodology is not meant to be prescriptive - steps may be done in a different order and rules of
thumb should be carefully evaluated by the designer and treatment plant operators. The key is that any
methodology be based on the fundamental science of biological nutrient removal as presented in
Chapters 4 and 5 of this manual. In addition to guidance in this section, the reader is also encouraged to
review steps and example calculations provided in the USEPA Nitrogen Control Manual (1993) and newly
updated Manual of Practice No. 8 by the Water Environment Federation (WEF) and the American
Society of Civil Engineers (ASCE), Design of Municipal Wastewater Treatment Plants (WEF and ASCE
2010).
Step 1: Characterize the influent flow.
See Chapter 7, Section 7.4 for detailed guidance. Key factors for design are as follows:
• The lowest mixed liquor temperature for determination of required SRT. Usually based on the
coldest temperature sustained for a two week period, but the period could be shorter if very
low SRTs are being considered.
Nutrient Control Design Manual 10-3 August 2010
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• The expected chemical oxygen demand (COD) to Total Kjeldahl Nitrogen (TKN) or biochemical
oxygen demand (BOD2) to TKN ratio, and the COD to Total Phosphorus (TP) ratio. The desired
ratio is that of the settled sewage if primary sedimentation is included in the treatment train.
For complete denitrification, a COD/TKN ratio of more than 8 and a BOD/TKN ratio of more than
4 are needed. For biological phosphorus removal, a COD/TP ratio of greater than 40 and
BOD/TKN ratio of more than 20 are needed when targeting effluent TP concentrations less than
1.0 mg/L A more accurate ratio for enhanced biological phosphorus removal (EBPR) is the
readily biodegradable COD (rbCOD) to TP ratio of the settled sewage (should be > 15, see
Chapter 5 for discussion). Note that if rbCOD is present in a truly anaerobic reactor (no
dissolved oxygen(DO), no nitrate or nitrite), some EBPR will always occur. The amount of soluble
phosphorus in the plant effluent will depend upon the rbCOD/TP ratio available to the
phosphorus storing (poly-P) bacteria in the anaerobic zone. The concentration of the poly-P
bacteria in the activated sludge shifts as the rbCOD/TP ratio shifts.
Step 2: Determine the required SRTfor nitrification.
Use graphs or equations provided in Chapter 4 to determine the minimum SRT for near
complete nitrification. Multiply by a safety factor - this is typically greater than 2 or represented by the
peak TKN divided by the median TKN in the process influent. Both of these approximations will provide
a large safety factor that can be reduced using dynamic simulation with a calibrated computer model.
Designers should carefully consider the effluent requirements including the required averaging period
when selecting a safety factor for preliminary design. For example, if the effluent limit is 3.0 mg/L, a
single value of 5.0 mg/L will be very difficult to average out if the averaging period is one month because
there is little possibility of producing an effluent concentration of less than 3.0 mg/L for most
wastewaters. However, if the averaging period is seasonal or annual, such an excursion is of significantly
less concern and a smaller safety factor should be used.
Step 3: Determine required SRTfor denitrification and phosphorus removal (if required).
Use graphs or equations provided in Chapter 4 to determine the denitrification rates at the
lowest design temperature and use that value to determine the SRT required for the anoxic zones. Add
the SRT requirements for denitrification to that required for nitrification. If biological phosphorus
removal is required, add an anaerobic volume of 1.5 to 2 hours. Recognize, however, that complete
characterization of the process influent wastewater and dynamic simulation might show that an
anaerobic nominal HRT period as low as 0.5 hours would be sufficient for high rbCOD/TP ratios entering
the plant, (for example, if the plant is fed by force mains and/or the wastewater is fully fermented when
it reaches the plant as is likely to happen for high temperatures). The anaerobic volume can be
significantly reduced by adding volatile fatty acids from pre-fermentation or by chemical addition for
polishing.
Step 4: Determine Overall Sludge Inventory.
A preliminary estimate of the sludge inventory for the BNR activated sludge system to be
designed can be derived from Figure 7-1, which is the same as Figure 14.20 from the WEF-ASCE Manual
of Practice 8 (2010). For the total SRT and temperature as identified in Steps 1 through 3, estimate the
2 For the purposes of this design manual "BOD" represents the 5-day BOD measurement method (sometimes
referred to as BOD5) unless otherwise noted.
Nutrient Control Design Manual 10-4 August 2010
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net sludge production as a function of influent BOD (i.e., assume that 100 percent of the BOD will be
removed during the treatment process). See Chapter 14 of MOP 8 (WEF and ASCE 2010) for guidance
on performing detailed calculations and determining the overall sludge inventory. Estimate the sludge
wastage rate as a function of the net sludge production and influent flow rate, i.e. calculate the mass of
BOD that will be removed per day. This requires assuming a nominal hydraulic retention time for the
total biological reactor volume. For conventional BNR activated sludge, this would be 6 to 10 hours
depending upon the design temperature and the MLVSS concentration desired. Using the factor
selected from Figure 7-1, calculate the mass of VSS that will be produced per day from the mass of BOD
that will be removed per day. Divide this number by the selected SRT to determine a rough
approximation of the waste activated sludge that will be produced per day.
Step 5: Estimate MLSS and Determine Overall Basin Volume.
Decide on a MLSS value, usually ranging from 2,200 to 3,500 mg/L if gravity sedimentation
clarifiers are being used. The MLSS concentration is based on the secondary clarifier design, if gravity
sedimentation, or on a reasonable concentration for membrane separation, e.g. 8,000 to 10,000 mg/L.
If Integrated Fixed Film Activated Sludge (IFAS) is used, designers need to determine the biomass that
will be attached to the media for the design operating conditions. That is best done using a computer
model. Final selection of MLSS may be done by trial and error and should consider solids settling and
thickening properties (WEF and ASCE 2010).
Multiply the SRT (days) from Steps 1 through 3 by the sludge wastage rate (mass /day) and
divide by the MLSS concentration (mass per unit of area) to estimate the required minimum reactor
volume.
Step 6: Determine Internal Recycle and RAS Pump Rates
Allow for mixed liquor internal recycle pumps that can return 2 to 4 times the influent flow for
nitrate recycle from the end of the first aerobic zone to the influent end of the anoxic zone. Also, if the
University of Capetown (UCT) or the Virginia Initiative Project (VIP) configurations are being used,
recycle MLSS at a rate of approximately 1 times the influent flow from the end of the anoxic zone to the
influent of the anaerobic zone for EBPR. The internal recycle to the anaerobic zone is not needed if the
RAS is returned to the anaerobic reactor influent. Note that the most efficient internal nitrate recycle
rate depends on the COD/TKN ratio entering the anoxic zone. Above this rate, there is insufficient BOD
to reduce the quantity of nitrates entering the zone and additional recycle is detrimental rather than
useful because internal recycle typically recycles DO as well as nitrates. There is rarely justification for
recycle rates greater than 4 times the influent flow because of the increase in pumping costs for the
same amount of additional nitrate that will be reduced, even if sufficient biodegradable COD is available.
The RAS rate should generally be 0.5 to 1.2 times the influent flow rate. The operator should be
provided with substantial RAS rate flexibility so that adjustments can be made for variations in sludge
settling rates and for changes in the MLSS concentration in the biological reactors.
Designers also should allow flexibility for when the plant is under loaded during the early years
of operation or when the organic loading is considerably less than the design loading. Because under
loaded BNR processes do not perform as effectively as fully loaded BNR processes, the use of multiple
trains is the best way to provide flexibility during the under-loaded years. Also note that nitrification
processes require minimum amounts of alkalinity to function optimally. The actual amounts of alkalinity
Nutrient Control Design Manual 10-5 August 2010
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in the wastewater should be determined and compared to the net quantity that will be consumed by
nitrification and phosphorus precipitating chemicals after correction for alkalinity that will be produced
during denitrification.
10.3 Overview of Recommended Approach for Plant Modeling
Several groups have published recommended protocols and guidelines for activated sludge
modeling (Petersen et al. 2002; Hulsbeek et al. 2002; Melcer et al. 2003; Shaw et al. 2007). Recognizing
the need for a consistent protocol, IWA established a task group on good modeling practice to prepare
guidelines for the use of activated sludge models. The task group hopes to finalize the guidance in 2010
(Rieger 2009).
To date, the task group has developed a Unified Protocol for modeling with the following five
key steps:
1. Project description, including identification of objectives and model requirements
2. Data collection and reconciliation
3. Model set up
4. Calibration and validation
5. Simulation and results interpretation
Figure 10-1 shows the Unified Protocol as published by IWA. Modelers should generally follow this
protocol when designing retrofits to an existing treatment plant to enhance nutrient removal.
Nutrient Control Design Manual 10-6 August 2010
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Agreement on the objectives and
budaet?
Data collection
(physical, operational, influent, effluent)
Process flow diagram (process units, boundaries)
Validation successful'
y
Calibration/validation adequate
Functional evaluation
Figure 10-1. Unified protocol for activated sludge modeling.
Source: IWA, 2009. Used with permission from the IWA (http://www.modeleau.org/GMP TG/UP.htm), from an upcoming report
(http://www.iwapublishing. com/template. cfm?name=isbn9781843391746&tvpe=forthcoming)
Nutrient Control Design Manual
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August 2010
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10.4 Establishing Objectives and Requirements
As early as possible in the modeling process, designers and plant operators should agree on the
following:
• Modeling objectives
• Boundaries, answering questions such as "Under what environmental and operating conditions
will the model be used?" and "What types of simulations will be performed?"
• Preconditions
• Performance criteria (e.g., goals for modeling accuracy)
• Responsibilities
• Required data
• Time and budget constraints
Three topics—intended use of modeling, goals for accuracy, and types of simulations—are particularly
important in the design of nutrient removal systems and are discussed in subsequent subsections.
10.4.1 Intended Use of Modeling
This design manual assumes that modeling will be used for the design of retrofits to existing
plants to enhance nitrogen and/or phosphorus removal, but can be used for green field designs as well
The same simulators developed for design purposes, however, can have many other functions.
Simulators can be used to assess unit process capacities under different operating conditions to identify
which processes are critical bottlenecks. They can be very useful as a process optimization tool through
the manipulation of various controllable parameters. The same simulator used for design can be used
to optimize performance once the retrofits are in place. Designers should also determine if the
simulator will be used for operator training—for example, to simulate "what if?" scenarios that could
have significant impacts on plant performance.
10.4.2 Goals for Modeling Accuracy
Because the simulator is a simplified representation and does not account for every biological
process and chemical reaction that occurs in the real world, a perfect calibration to observed
performance is not practically achievable and is not a recommended standard of modeling. An
acceptable level of agreement is generally for observed and predicted values to be within 10 to 15
percent for an initial design when calculated on a monthly average basis and below 10 percent for a final
design. With on-line instrumentation that can track the diurnal variations, it may be possible to
approach 5 percent accuracy on a daily average basis for model calibration of most of the parameters
such as sludge production, MLSS, effluent ammonia and soluble COD. Some effluent parameters such as
nitrates are dependent on the degree of simultaneous nitrification and denitrification, which is more
difficult to model. Agreement will be less for transient conditions further away from steady state such as
storm events or spikes in nutrient loading. The goal for modeling accuracy should take into account the
design criteria for nutrient limits. A smaller margin of error is preferable for very low target effluent
levels for TN and/or TP.
Designers should coordinate closely with the plant owner to establish target accuracy prior to
developing the model. Note that the model output is only as accurate as the input and the method
Nutrient Control Design Manual 10-8 August 2010
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used to make the measurement for inputs and outputs. The better the input information, the lower the
safety factor that is needed.
10.4.3 Dynamic vs. Steady State Simulation
Models can be run in steady state or dynamic simulation mode. Steady state simulations predict
treatment plant performance when there is no variation in the load or flow to the plant. Diurnal
patterns for influent flow volume and loadings (i.e., COD and TKN) may be used in steady state models,
but the models will average the flow and load and produce steady state outputs. Dynamic models rely
on the same mathematical relationships as steady-state models, but they can also simulate hourly
variations in measured influent flow and operating parameters (e.g., recycle flows) over a number of
days, months, or seasons.
Industry experts consider dynamic modeling to be superior to steady state modeling (Hulsbeek
et al. 2002; Merlo et al. 2008; Alex et al. 2007). Dynamic simulations more closely approximate actual
plant operations and allow for the use of smaller safety factors for design. They are extremely useful for
wet weather modeling, determining daily maximum limits, and sizing aeration equipment for peak
demand (Johnson 2009). Design objectives, however, should be balanced with significantly larger data
requirements (e.g., hourly measurements of influent flow and nutrient loadings for several days) and
hence, the higher costs associated with building a dynamic compared to a steady state model. As noted
previously, however, the cost of more sampling and analysis is a fraction of the savings possible through
a more accurate design and lower safety factors.
Dynamic modeling becomes particularly important and is recommended when targeting very
low nutrient limits such as 3.0 mg/L for total nitrogen (TN) and/or 0.1 mg/L for total phosphorus (TP).
Reasons for this are discussed below.
Because nitrifying bacteria do not have adsorptive capacity, the rate at which they can reduce
ammonia is directly related to the mass of nitrifiers in the mixed liquor, which is in turn determined by
the average influent ammonia load and the SRT in the aerated basin. Influent ammonia concentrations
vary, with peaks occurring during the day and much lower concentrations occurring overnight. In fact,
peak ammonia load during the morning can be as high as 2.8 times the average (Barnard 1975). Influent
ammonia concentrations can be greatly increased by concentrated ammonia in return streams from
sludge dewatering operations. In many plants, dewatering of digested sludge occurs during the day,
and the supernatant from the dewatering operation is returned to the head of the plant in the
afternoon. The combination of peak ammonia loads from the influent wastewater and additional
ammonia from return streams can result in peaks that cannot be fully oxidized by the available nitrifiers
resulting in peaks of ammonia in the effluent. The size of the peaks will also be related to the SRT and is
not necessarily an indication of imminent plant failure, i.e. nitrifier washout. Steady state modeling
would not capture this scenario. When there are substantial daily ammonia peaks, the steady state
model may show that nitrification is possible while the diurnal model may indicate inadequate
performance of the nitrifiers because of ammonia breakthroughs. For example, when the average
influent nitrogen is 30 mg/L and there is an increase to 60 mg/L without an increase in the SRT during
the transition from 30 to 60 mg/L, some ammonia will bleed through until the mass of nitrifiers has
doubled. In simulation or actual practice, this would take a significant number of days. It is unlikely that
the high load will persist long enough for the increase in nitrifiers to take place unless there has been a
permanent addition to the ammonia load. Therefore, significant variations in the effluent ammonia
concentration could be observed but there is little or no danger of losing nitrification. Note also that
Nutrient Control Design Manual 10-9 August 2010
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since the nitrifier population is a low fraction of the total MLSS, the total MLSS will increase only slightly
(2 to 3 percent) if a permanent increase in ammonia loading has occurred.
Similar to the reduced treatment capacity of nitrification systems caused by ammonia in recycle
from sludge dewatering units, reduced phosphorus removal can be caused by nitrates in the recycle
stream. This is because some heterotrophic aerobic organisms, including some phosphorus storing
organisms, can use nitrate as an electron acceptor and will compete with the phosphate accumulating
organisms (PAOs) for volatile fatty acids (VFAs). See Chapter 5 for additional information on the
biological phosphorus removal process and inhibitory effects of nitrate in the recycle stream.
The main simulator models incorporate a one-dimensional (ID) mathematical model for
secondary clarifiers based on the mixed liquor settling characteristics as represented by the Vesilind
equations (Vesilind 1968; Zhou et al. 1982). These ID models can predict the behavior of transient loads
through the secondary clarifiers during dynamic modeling, which is extremely important when removing
nutrients to low levels. The ID models, however, are limited because they assume ideal layered
clarification and thickening. Two or three dimensional (2D or 3D) computational fluid dynamic (CFD)
models can more accurately represent the dynamic behavior of specific clarifier configurations (Merlo et
al. 2007). CFD models could be coupled with the overall simulator models; however, they require a very
large computer memory to run, which slows down the iterations required to solve the multiple
interrelated biological equations. Most designers use the ID models for solving the biological reactions
and the effect of transient loads, then study the final clarifiers using CFD modeling.
If designers would like to see the effect of dynamic modeling but only daily average and some
peak flow values are available, they can use typical patterns for flow and concentration derived from
similar plants (it should be noted on the outputs that these are estimates). Designers can use a separate
input of clean water to simulate a storm event of fixed duration then mix this input with regular plant
influent. The storm could be simulated at various times during a 24-hour period to assess its potential
impact. Recycle flows can also be simulated as being stored and fed at various times within a 24-hour
period under various stress conditions, such as low temperature or low VFAs in the plant influent. For
example, the centrate from digested sludge could be modeled for return over 8 hours, 24 hours, or to be
discharged during the night when the incoming ammonia load is low. Some plants may run out of
alkalinity required for nitrification at low flows and some programs can simulate this.
10.5 Selecting a Process Simulator
Process simulators allow users to link unit treatment processes such as bioreactors and clarifiers
together according to the configuration of a particular plant. Each unit process incorporates a
mathematical model to represent biological, physical, and/or chemical interactions within that unit. The
process simulator uses this information to mimic the performance of a plant under specific operating
and influent loading conditions. Each individual biological reactor in the models is assumed to be
completely mixed.
Table 10-1 lists the commonly used commercially available simulators (this is not an exhaustive
list as the simulator market is continually changing). Each simulator is COD-based, allows dynamic and
steady state simulations, has graphical interface and graphical output, and computes true mass
balances. Each simulator can also perform "whole plant" modeling, which includes not only primary
treatment processes, but also solids treatment and recycle streams. Commercially available simulators
vary, however, according to their graphic interface, the mathematical models included for specific
Nutrient Control Design Manual 10-10 August 2010
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treatment processes, the degree of customization available, the degree of knowledge required, and
cost.
Table 10-1. Commonly Used Process Simulators
Simulator
Biowin
GPS-X
SI MBA
STOAT
WEST
AQUIFAS
Distributor
to U.S.
Evirosim
Associates Ltd.
(Flamborough,
Ontario)
Hydromantis
Inc. (Hamilton,
Ontario)
ifake. V
(Germany)
WRc pic
(Swindon,
England)
DHI Water
(H0rsholm,
Denmark)
Aquaregen
Mountain
View, CA
USA
Description
Uses a proprietary combined model that
is based on the ASM models but
integrates them with models for fixed
film, digestion, settling, chemical
addition, and filtration ("supermodel"
approach). It is customizable and
compatible with Excel, and has a unique
pH calculator that can determine pH
changes caused by wastewater reactions
and model precipitation of struvite and
other minerals.
Uses a proprietary model based on the
ASM models. Includes models for fixed
film, digestion, settling, chemical
addition, and filtration. It is customizable
and compatible with Excel and Matlab. It
is able to perform automatic sensitivity
analyses and has an extensive process
model library.
Based on the ASM models, also has
models for settling and digestion. It is
customizable and uses Matlab as a basis.
Well designed to look at control
elements, it can link with collection
system and stream models.
Allows use of modified version of ASM
models that are based on BOD instead of
COD. It is customizable and can perform
automatic sensitivity analyses. UK specific
data.
"Open" system, i.e., can use any
mathematical model to predict plant
performance.
An internet available program based on
the ASM models that utilizes Excel. It
includes models for the incorporation of
all types of fixed-film media into activated
sludge bioreactors. Also MBR and whole
plant simulation.
Use
Along with GPS-X,
dominant in North
America
Along with BioWin,
dominant in North
America
Mainly Germany
and Holland
Predominantly
United Kingdom
Belgium and
Europe
Developed in the
USA, but currently
used in several
countries around
the world.
Website
http://www.envirosim
.com
http://www.hvdroma
ntis.com
http://simba.ifak-
md.de/simba/index. p
hp?option=com cont
ent&task=view&id=34
<emid=104&lang=e
n
http://www.wrcplc.co
.uk/default.aspx?item
=1052
http://www.dhigroup.
com/Solutions/Solutio
nSoftware/WEST.aspx
http://www. aquifas
.com
Source: Melcer et al. 2003; Shaw 2009. Sen and Randall 2008a,b,&c. Weblinks current as of March 23, 2010.
Two simulators, BioWin and GPS-X, dominate the market in North America. Both are relatively
easy to use and produce output data in simple graphical format (Shaw et al. 2007). In 1995, the Dutch
Foundation of Applied Water Research (STOWA) recommended the use of SIMBA, which prompted its
adoption by more than 100 plants in the Netherlands (Hulsbeek et al. 2002). STOAT, on the other hand,
Nutrient Control Design Manual
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August 2010
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is the only simulator that includes biochemical oxygen demand (BOD)-based models. West is popular
with some in the academic community because of its ability to accommodate any model and because it
runs in a transparent environment.
New simulators are continually being developed, some to meet specific needs of certain users.
For example, AQUASIM and ASIM are produced by EAWAG and have found high usage among
academics (for more information on these models, see http://www.asim.eawag.ch/). Sen and Randall
(2008a, 2008b, and 2008c) developed an internet available spreadsheet model called "Aquifas"
primarily for modeling attached growth processes including membrane bioreactors (MBR), integrated
fixed film activated sludge (IFAS) systems, and moving bed bioreactors (MBBR), but it can be used for
whole plant simulations of all types of activated sludge systems, including headworks, anaerobic
digesters and sludge processing, and is adaptable for fixed film processes such as trickling filters and
RBCs. Users have the option of selecting empirical models or biofilm diffusion models for predicting
performance and can run both steady state and dynamic simulations. Aquifas is available free of charge
at http://www.aquifas.com/.
While no set of mathematical equations can perfectly replicate a system, designers should be
careful to select a simulator that contains all significant treatment processes occurring in the plant and
candidate processes under consideration for upgrades. The choice of simulators will also depend on
user preference and skill level. While some simulators allow a great deal of flexibility to adjust and
change models, they also require greater skill and understanding of the process. The graphical interface,
user friendliness, and compatibility with existing software will also be important factors. Some
programs have a high initial cost but will most likely offer training and a higher degree of customer
support, while others cost less but do not provide as much support. The text box on the next page
provides a list of questions that should be considered by designers prior to selecting a process
simulator.
Nutrient Control Design Manual 10-12 August 2010
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What questions should I ask before selecting a process simulator?
• How will you use the simulator? What do you want the simulator to do now and in the future?
- Decide on FEATURES.
• Who will run the simulations and keep the models up to date?
Determine expertise needed to run and maintain the simulator.
• Questions for Vendors:
- What processes and functions will it model?
How does it interface with other software such as Excel?
Is there a lease option? This type of agreement can be very cost effective.
- What is the cost and availability of training?
- What type of support (from company and/or user groups) is offered?
How extensive are the documentation and help functions?
Source: Shaw 2009.
10.6 Data Collection
Figure 10-2, adapted from Wilson and Dold (1998), shows the inputs needed to set up a process
simulator for an existing WWTP. Input data that can be obtained from plant records or by direct
measurement include:
• Process configuration data such as plant layout, reactor sizing, and recycle streams
• Process operating conditions such as flows through each process and flow rates of recycle
streams
• Process inputs, namely wastewater characteristics and influent loading data
Rieger (2009) estimates that 50 to 60 percent of the entire modeling effort goes into data collection and
reconciliation.
Subsections 10.6.1 and 10.6.2 discuss process configuration and operating conditions,
respectively. Guidance for wastewater characterization and influent loading data is provided in Chapter
7. For guidance on estimating fractions of organic substrate and nutrient loading parameters, see
Sections 10.7 and 10.8, respectively. Estimating kinetic and stoichiometric parameters for process
models are discussed in Section 10.9.
Nutrient Control Design Manual 10-13 August 2010
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PROCESS MODELS:
Biological
(aerobic, anoxlc,
anaerobic reactions)
Physical-Chemical
PROCESS INPUTS:
Wastewater Characteristics
Loadings
Dynamic Patterns
TREATED
EFFLUENT
BIO-
REACTOR
SECONDARY
SF.TTL PR
SOLIDS
PROCESSING
RESIDUAL
SOLIDS
PROCESS
CONFIGURATION:
Flow Routing
Unit Sizes
Reactor Staging
Recycle Stream
PROCESS OPERATING
CONDITIONS:
Recycle Rates
Control Setpoints
Figure 10-2. Essential requirements for wastewater treatment process simulation.
Source: Wilson and Dold 1998. Reprinted with permission from Proceedings of WEFTEC®. 1998, the 71st Annual Water
Environment Federation Technical Exhibition and Conference, Orlando, Florida, October 3-7,1998. Copyright © 1998, Water
Environment Federation, Alexandria, Virginia, www.wef.org.
The accuracy of any simulator is highly dependent on the quality of the data used to build it.
Designers should critically evaluate data to ensure that they are logical (i.e., data are not outside the
normal range for WWTPs) and measured parameters are as expected for each sampling location given
the upstream treatment processes. Chapter 7', Section 7.4.2 recommends data checks that should be
completed prior to entering information into the model. Mass balances to further validate influent
loading and plant operating data are discussed in Section 10.10 (model calibration).
10.6.1 Process Configuration
Simulators require users to configure unit processes as they occur at the WWTP and enter
information on the size of each component, points of input and output, and recycle streams such as the
RAS and MLSS recycle. Most simulators allow users to "build" a plant by selecting icons for unit
processes and entering basic data for each (e.g., volume, length, width, and depth for aeration basins
and clarifiers). The end result is a visual representation of the existing WWTP to be upgraded or
retrofitted or the concept for the Greenfield plant. See Figure 10-3 for an example of a simulator
configuration for a Three-stage Phoredox Process (A2O) plant with primary sedimentation tanks, a
fermenter for VFA generation, waste activated sludge thickening, digestion, and dewatering. Designers
should always consider "whole plant modeling" for nutrient removal processes, meaning that sludge
production and recycle streams are configured in the simulator rather than just the major treatment
processes.
Nutrient Control Design Manual
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/August 2010
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Influent Mixer PST
Anaerobic
Anoxic Aerobic Aerobic
Clarifie
Figure 10-3. Example simulator configuration for a biological nutrient removal plant (courtesy of EnviroSim Associates, LTD).
Nutrient Control Design Manual
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All the major processes including secondary clarifiers, bioreactors, and any filters should be
included. If an objective of modeling is to evaluate aeration system performance, diffuser dimensions
and densities should also be included (Melcer et al. 2003). It is not usually necessary to include sludge
treatment process (e.g., digestion), although any return streams from these processes should be
specified. If possible, parallel treatment processes should be modeled as individual units (Shaw et al.
2007).
Information on physical configuration can be obtained from plant managers or operators or can
be obtained from as-built construction records. During initial model setup, designers should compare
design capacities for each unit to the modeled outputs during maximum operating capacity (Third et al.
2006). See Section 10.9 for additional information on model runs and calibration.
The set up of suspended growth bioreactors is very important for accurate modeling of carbon
removal and nitrification. The simulators use continuous stirred tank reactors (CSTRs) to represent the
suspended growth activated sludge systems, meaning that a drop of water is instantaneously mixed
with the rest of the water as soon as it enters that reactor. In real plants, the flow through bioreactors is
somewhere between plug flow and completely mixed depending on design and operational factors such
as shape of the reactor, flow rate, baffling, and mixing pattern of aeration equipment. To account for
non-ideal mixing conditions, individual bioreactors should be represented by at least two CSTRs in
series for square tanks and more than two for rectangular tanks in the model.
If the design of the reactor favors plug flow conditions (e.g., long narrow tanks with no dividing
wall between cells), designers can use the following formula to estimate the number of CSTRs in series
for the model:
N = 7.4* L*Q(1 + RR) Eq. 10-1
W* H
Where:
N = equivalent number of tanks-in-series (should never be greater than 12)
L = aeration tank length (m)
Q = wastewater flow (m3/sec)
RR = RAS recycle ratio (dimensionless)
W = aeration tank width (m)
H = water depth (m)
This formula was developed by the Water Research Centre in the United Kingdom. For additional
information, see Melcer et al. (2003).
Another more rigorous approach to characterizing flow through a bioreactor is to perform a
tracer test. A slug of inert tracer (e.g., lithium or fluorescent dye) is added to the head of the tank and
monitored to develop a washout curve. See the publication Guidance Manual for Sewage Treatment
Plant Liquid Train Process Audits (Environment Canada 1995) for the detailed methodology.
The main reason for matching the number of individual completely mixed tanks with the shape
of the existing or proposed tank is to simulate the different operating conditions in different zones. E.g.
if the length to width ratio for the aeration section is 3, the oxygen demand in the first section will be
Nutrient Control Design Manual 10-16 August 2010
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higher while the dissolved oxygen level in this section must at least 2 mg/L to minimize bulking sludge
development and to enhance both nitrification and biological phosphorus uptake. In contrast, by the
time the mixed liquor reaches the third section, the ammonia may be depleted resulting in a drop in
oxygen demand which may lead to over-aeration. Air flows to the different zones should match the
demand to save energy and ensure that the biological functions work optimally. In addition, there may
be a mixed liquor recycle from the aeration basin to the anoxic zone which would change the flow
patterns in the aeration basin
10.6.2 Operating Conditions
The following operating parameters are important model inputs and are available at most
WWTPs:
• Flow routing into and out of major unit treatment processes
• RAS and mixed liquor recycle rates
• Primary and secondary wastage flows (with schedules in cases of intermittent wasting)
• Flow rate and schedule for return streams that contain high solids, nitrogen, or phosphorus
• Control schemes for mechanical equipment such as aerators and pumps (based on SCADA
system records or interviews with plant operators)
Designers should review several years of data to determine operating conditions under dry weather,
normal operation rather than taking an average of several months of data. Operating parameters
should also be established for extreme conditions such as peak wet weather events.
Plants may be missing operating information, such as the proportion of flow going to multiple
treatment trains. Designers should interview operators to approximate missing operational data or use
portable equipment (e.g., flow meters) to directly measure and record missing parameters. Site visits
to the treatment plant to interview operators, conduct a walk-through of the plant, and observe
treatment processes are highly recommended (Hulsbeek et al. 2002).
A common suspect operating parameter is wastage flow due to the difficulty in measuring
sludge densities over the full period of wastage (Melcer et al. 2003). See Section 10.10 for guidance on
checking wastage flow rates during calibration.
10.7 Characterization of Organic Material
When using models for design, organic compounds need to be differentiated into a number of
separate components based on their characteristics, namely whether the material is soluble3 and/or
biodegradable. The relative fractions of organic material in wastewater are very important because they
determine the amount of substrate available to microorganisms for biological nutrient removal
processes and affect the amount of DO required and the volume of sludge produced. Most models
3 Melcer et al. (2003) reports a lack of clarity in use of the term "soluble." Most technical references define soluble
as material passing through a 0.45 micron (urn) membrane filter. Total dissolved solids (TDS), however, are
routinely measured using glass-fiber filters with nominal pore sizes of 1.2 to 1.5 u.m. The difference is in the
colloidal material, with most colloidal COD passing through a glass-fiber filter but retained on a 0.45 urn filter. It is
therefore very important to note the filter type used when examining historical data and document the filter type
used for all experiments.
Nutrient Control Design Manual 10-17 August 2010
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include default values for organic fractions; however, site-specific characterization is strongly
recommended for the design of nutrient removal systems.
Characterization of organic material is almost always in terms of COD rather than BOD4. Even if
the simulators allow users to enter BOD values, they are converted to COD for mass balances or
simulations, using the COD fractions in the model. For additional rationale, see the text box in Chapter
7, Section 7.4.1. When COD data are not available, COD can be estimated based on its relationship to
measured BOD and VSS values (see Section 10.7.1 for guidelines). It is also common when only BOD
values are available to run a series of BOD and COD determinations on the same samples to find a
reasonable correlation curve which can be used to determine the COD for every BOD value.
Figure 10-4 provides a schematic representation of the COD components in wastewater.
Commonly used symbols for each component along with a description and overview of measurement
methods are provided in Table 10-2. The naming convention for the symbols is "S" for soluble, "X" for
particulate, "I" for unbiodegradable or inert. COD components are typically reported in milligrams per
liter (mg/L), but can also be represented as fractions of the total as indicated by the letter "F" with the
symbol as a subscript (e.g., FS! = the fraction of total COD that is soluble and unbiodegradable).
As shown in Figure 10-4, total COD in the plant influent is the sum of the soluble biodegradable
portion, including both volatile fatty acids (VFAs) and non-VFA compounds; the soluble unbiodegradable
(or "inert") material; slowly biodegradable material, which can be either colloidal or particulate; and
particulate unbiodegradable material. As a formula, this relationship is expressed as:
COD-riNF = SA + SF + S: + SCOi_ + Xs + X: Eq. 10-2
Where:
CODT|NF = Total influent COD concentration
SA = Volatile fatty acids
SF = Complex biodegradable soluble COD
S: = Soluble unbiodegradable COD
SCOL = Slowly biodegradable colloidal COD
Xs = Slowly biodegradable particulate COD
X: = Particulate unbiodegradable COD
4 The STOAT simulator contains modified mathematical expressions that allow the user to characterize wastewater
using BOD instead of COD. STOAT is the only one of the commonly used process simulators shown in Table 10-1
with this feature.
Nutrient Control Design Manual 10-18 August 2010
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Readily
Biodegradable
(Ss or rbCOD)
Slowly
Biodegradable
(SBCOD)
Soluble VFA (SA)
Soluble Non-VFA
(Ss)
Soluble
Unbiodegradable (S:)
Slowly
Biodegradable
Colloidal (SCOL)
Slowly
Biodegradable
Particulate (Xs)
Particulate
Unbiodegradable
Flocculated
"and Filtered
COD
^"Filtered"
COD
Total
COD
COD
Figure 10-4. COD components for municipal wastewater
Note: COD fractions are not to scale and vary from plant to plant.
Source: Derived from Melcer, H., P.L. Dold, R.M. Jones, C.M. Bye, I. Takacs, H.D. Stensel, A.W. Wilson, P. Sun, and S. Bury.
2003. Figure 4-5, with Permission from the Water Environment Research Foundation.
Nutrient Control Design Manual
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Table 10-2. COD and Particulate Fractions in Municipal Wastewater
Fraction
Total Influent COD
Readily
Biodegradable
COD
Volatile Fatty
Acids
Complex
Biodegradable
Soluble COD
Soluble
Unbiodegradable
COD
Slowly
Biodegradable
COD
Slowly
Biodegradable
colloidal COD
Slowly
Biodegradable
Particulate COD
Particulate
Unbiodegradable
COD
Symbol1
CODT,INF
Ss(or
rbCOD)
SA(or
VFAs)
SF
Si
SBCOD
SCOL
Xs
Xi
Description
Quantifies "strength" of
organic material in the
influent
Can be easily absorbed by
organisms and used for
energy and synthesis of
cell mass. Is the sum of SA
and SF
A fraction of Ss
The fraction of Ss that is
not VFA
Portion of soluble COD
unaffected by biological
reactions at the plant.
Leaves the secondary
clarifier at same
concentration as influent
Portion of biodegradable
COD that requires
extracellular enzymatic
breakdown prior to
adsorption and utilization.
Is the sum of SCOL and Xs
Portion of SBCOD that is
colloidal and typically not
settleable
Potion of SBCOD that is
particulate and settleable
Portion of particulate COD
unaffected by biological
reactions at the plant.
Accumulates in sludge
mass.
How It Can Be
Measured
Directly measured or
estimated based on
relationship to BOD
Directly measured by
respirometry, but other
methods are available
using simplifying
assumptions
Directly measured using
ion or gas chromatography
Ss-SA
Approximated as the
soluble (filtered) COD of a
well nitrified plant effluent
Typically determined as
CODrjNF-Si-Ss-Xi
The difference between
the filtered COD and the
ffCOD of the effluent
SBCOD - SCOL
Determined from the
model or estimated based
on influent COD, BOD, and
TSS
Concentration in
Municipal
Wastewater2
(mg/L)
250 - 700
25 - 125
20-50
200 - 400
35 - 110
Notes:
1. The literature contains more than one symbol for some components. The symbols shown are generally consistent with
Melcer et al. (2003). Other commonly used symbols are included in parentheses. Note that the symbol shown represents
concentration, expressed as milligrams per liter (mg/L). Fractions of total COD are represented by the letter "F" and a subscript
(e.g. FS| is the fraction of the influent COD that is Unbiodegradable and soluble).
2. Derived from Melcer et al. (2003), Table 4-2 and experience with systems. Concentration may vary due to variable per capita
water consumption.
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The significance of the COD fractions on wastewater treatment is described below.
• Readily biodegradable COD (Ss or rbCOD) is an extremely important modeling variable for
biological phosphorus removal (BPR) systems as it determines the amount of substrate available
for PAOs to use in the anaerobic zone. It includes VFAs and soluble, readily biodegradable COD
that are not VFAs but can degrade to VFAs in the anaerobic zone.
• Slowly biodegradable COD (SBCOD) is made up of particulate and colloidal material that
requires extracellular breakdown prior to adsorption by microorganisms. The difference
between the particulate fraction (Xs) and the colloidal fraction (XCOL) is important when
modeling performance of primary clarifiers because a portion of the particulate fraction will
settle and be removed with the waste sludge while all the colloidal matter will pass through to
the aerobic process.
• Soluble unbiodegradable COD (Si) is significant in that it passes through the plant and ends up
in the plant effluent. It is important to distinguish this portion from the soluble biodegradable
fraction in the plant influent. S: is typically a small fraction of the total.
• Particulate unbiodegradable COD (Xi) will accumulate in the waste sludge. From a mass
balance perspective, the mass of X: in the system is equal to the mass coming into the system
per day multiplied by the SRT. X: accounts for increasing VSS in the sludge with increasing SRT.
In addition to the organic matter fractions described above, active biomass will be captured by
total COD measurement methods and is reflected in particulate COD fractions (X: and Xs). Orhon and
Cokgor (1997) estimate that active biomass makes up between 7 and 25 percent of total COD in
municipal wastewater. Active biomass COD is not typically measured or taken into account for modeling
and will not significantly impact model predictions of plant performance except in some high rate (low
hydraulic residence time(HRT)) systems. See Melcer et al. (2003) for determining active biomass COD if
you believe that it might be a significant fraction in your system.
The aggregations on the right side of the schematic in Figure 10-4 relate measurement methods
to the representative fractions. The filtered COD is measured after the sample has been filtered through
a 0.45 u.m filter and represents both very fine colloidal and soluble material. The colloidal material can
be removed by flocculating the sample before filtering it. The result, called the flocculated and filtered
COD or ffCOD, represents only the readily biodegradable and truly soluble unbiodegradable COD.
Important relationships between organic and particulate material are discussed in Section
10.7.1. Section 10.7.2 presents various methods for determining COD fractions for a given wastewater.
Recommended data checks for COD fractions are provided in Section 10.7.3.
10.7.1 Relationship of Organic Material and Suspended Solids in Wastewater
Figure 10-5 shows the relationships between BOD, COD, TSS, and VSS along with key fractions of
each. Understanding these relationships can help designers estimate COD fractions when data are not
available or check results against other parameters when COD fraction data are available.
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BOD
COD
Figure 10-5. Relationship between BOD, COD, TSS, and VSS.
vss
TSS
Note: BOD = biochemical oxygen demand; COD = chemical oxygen demand; TSS = total suspended solids; VSS = volatile
suspended solids; X: = particulate unbiodegradable COD; S: = soluble unbiodegradable COD; VFA = volatile fatty acids; ISS =
Inorganic suspended solids
BOD can be either soluble or particulate. The soluble BOD is represented in COD as the total of
the readily biodegradable portion (Ss) and the colloidal portion of SBCOD. The particulate slowly
biodegradable COD corresponds to the particulate fraction of BOD. The particulate and soluble
unbiodegradable fractions of COD are not captured by the BOD5 method. The COD/BOD ratio for most
municipal wastewater is 1.9 to 2.2 mg COD per mg BOD (see the text box in this section for guidance on
estimating COD by determining the COD/BOD ratio).
Total suspended solids (TSS) represent the particulate portion of COD plus the inorganic
suspended solids. TSS is typically determined by weighing the solids retained on a 1.2 u.m glass fiber
filter. The sample can then be ignited and weighed again. The weight retained is the inert suspended
solids (ISS), and the difference between TSS and ISS is the volatile suspended solids (VSS). The ISS is an
important parameter in modeling. These solids pass through the plant unreacted and are removed in
the waste sludge, with a small portion leaving in the effluent. ISS can have a significant effect on the
mixed liquor suspended solids. In plants with no phosphorus removal or chemical addition, a plant
balance on ISS is a good check on the solids fractions as the ISS into the plant should equal the ISS out.
When phosphorus is removed from the liquid phase by either biological or chemical means, it leaves the
plant through the solids and registers as ISS.
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The particulate fraction of COD can be correlated to historical data on VSS of the mixed liquor
(MLVSS). This ratio is typically between 1.42 and 1.48 mg COD per mg MLVSS; however, the ratio can be
higher for solids other than biomass. Some models use a constant COD/VSS ratio to determine VSS
throughout the plant, while others may allow different values for different solids fractions such as
biomass, X:, and Xs.
What if the plant does not have COD data? This is a common problem. Many plants measure BOD5
instead of COD because the BOD method has been in use for a long time and it is often the basis for
regulatory compliance. If COD measurements are not available, COD can be approximated by
establishing the ratio of COD/BOD or COD/VSS. While some programs may take BOD as the input, it is
still necessary to adjust the COD fractions for good correlation.
Although measured values of BOD, COD, and VSS in the plant influent vary greatly, the ratios are
generally consistent for a specific wastewater. COD/BOD ratios are typically 1.9 to 2.2, with the higher
ratios for relatively fresh wastewater. COD/ VSS should range from 1.42 to 1.48 mg COD per mg
MLVSS. Estimating the ratio of COD/BOD is more common than COD/VSS.
To determine the COD:BOD ratio for a specific wastewater:
• Collect several samples (10 to 20 are recommended) of influent wastewater. The samples
should be representative of total influent flow and taken at different times of the day at low
and high influent flow rate.
• Split the samples, taking care to homogenize them thoroughly.
• Analyze each sample for COD and BOD. Use the same method for BOD that is used for the
available plant data. (BOD5 is most common, although some plants will have cBOD5 data. For
most municipal wastewaters there is very little difference in BOD5 and cBOD5 if settled primary
sludge is used as the seed because very few nitrifiers will be present and they grow too slowly
to significantly affect the results after just five days. If, on the other hand, effluent from a
nitrifying plant is used as the seed, the results will be substantially influenced by nitrification
and the answer will misrepresent the biodegradable organic matter in the wastewater. )
• Determine the average COD:BOD ratio for all samples.
• Repeat the test with primary effluent
Designers can then apply this ratio to historical BOD data to determine COD.
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10.7.2 Methods for Determining COD Fractions
Industry literature provides many different approaches for site-specific characterization of COD
fractions. This section presents the following three possible approaches, moving from the easiest to the
most complex:
1. Conduct a mass balance of the COD/ VSS ratio and BOD around the primary clarifier using model
default values for COD fractions. This method requires minimal sampling and can be a good first
screening step based on existing data.
2. Conduct special sampling of the plant influent and effluent to estimate each fraction. Several
measurement methods and approaches for approximating fractions are available. At minimum,
this method requires a dataset for filtered COD and flocculated and filtered COD (ffCOD) for the
plant influent and effluent. Advanced measurement techniques such as respirometry are
recommended in some cases.
3. Bench scale sequencing batch reactor (SBR) tests (also called the Low F/M SBR test). This
approach requires a 6-week start-up period and 2 weeks of intensive monitoring (Reiger 2009).
It is the most expensive (approximately $50,000) but can provide excellent characterization of
COD fractions, TKN fractions, and nitrifier growth rate.
METHOD 1: Mass Balance around the Primary Clarifier
Because a large portion of the VSS will be removed in the primary sludge and the soluble COD
passes into the effluent, a mass balance around the primary clarifier can provide significant information
on COD fractions. The COD removed in the primary clarifiers should correlate with the removal of BOD
and VSS since the latter also registers as COD.
The mass balance requires the following information:
• COD and BOD data for the clarifier influent and effluent.
• TSS and VSS data for the clarifier influent and effluent. If either TSS or VSS data are not
available, they can be approximated by assuming a VSS/TSS ratio of 0.75 for clarifier influent and
0.83 for clarifier effluent.
• Flow rate into and out of the clarifier.
• Wastage rate.
• Percent solids removal.
Enter these data into the model and use default values for COD fractions. Run a mass balance around
the primary clarifier. If the mass balance closes (i.e., the effluent BOD calculated in the model
approximate the actual BOD effluent) and the predicted effluent COD and TSS values match the actual
values, the default values for COD fractions are good approximations. If the mass balance does not
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close or the predicted and actual values do not match, then the default values for the COD fractions
need to be adjusted. See Methods 2 and 3.
METHOD 2: Special Sampling of Plant Influent and Effluent
This section presents recommended methods for estimating each COD fraction. In many cases,
more than one method is available.
Readily Biodegradable COD (Ss)
The simplest way to determine Ss is to measure flocculated and filtered COD (ffCOD) of the
influent and effluent of a plant producing low effluent ammonia and BOD, and assume the following:
• All of the readily biodegradable COD (S:) is consumed in the plant and negligible in the plant
effluent.
• There is no generation of soluble unbiodegradable COD (S:) in the plant (i.e., it is the same
concentration in the influent and effluent).
Ss is then determined as the difference between the influent and the effluent ffCOD.
The assumption that effluent Ss is zero generally holds when the sludge age is 3 days or older.
The assumption regarding the generation of S: can be checked by measuring both the influent and
effluent. If either assumption is not valid or if results are not as expected, Sscan be determined using
respirometric techniques. Several techniques are available and can be found in Melcer et al. (2003).
To distinguish between the VFA portion of Ss, measure VFA in the plant influent directly using
ion or gas chromatography.
Soluble Unbiodegradable COD (Sj
S: can be determined based on the following assumptions:
• There is no generation of S: within the system. While it is possible that some microbial
processes may add S:, this is generally a small portion of the total.
• The soluble biodegradable COD in the plant effluent is negligible compared to the
unbiodegradable portion. This is particularly true if the effluent ammonia is less than 0.2 mg/L.
The effluent soluble biodegradable COD is assumed to be zero.
• The effluent has very little colloidal matter because it is absorbed on mixed liquor solids (this is
generally the case for municipal wastewaters but may not hold true for industrial wastes).
Therefore, effluent colloidal COD is assumed to be zero. If colloidal matter is present, it can be
flocculated and filtered out.
It follows that S: can be approximated by measuring the soluble COD (i.e., the portion passing through a
0.45 u.m filter) in the plant effluent.
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Slowly Biodegradable COD (SBCOD)
SBCOD includes both particulate and colloidal COD and can generally be found by subtracting
the other portions of the COD from the total using the following equation:
SBCOD = COD-rjNF - S, - Ss - X, Eq. 10-3
To distinguish between the colloidal and particulate fractions of SBCOD, designers can compare
the filtered COD results to ffCOD results for influent samples. The difference is the slowly biodegradable
colloidal COD (Xs)
Particulate Unbiodegradable COD (X/J
Several methods are available for estimating X:. Melcer et al. (2003) reports that it is most easily
estimated by an iterative approach where X: is adjusted in the simulator until the predicted MLVSS
matches the observed values. Changing X: will also change the oxygen uptake rate (OUR), so OUR should
be checked to make sure it does not exceed the limits of the aeration system.
The X: fraction can also be estimated using historical plant data for COD, BOD, and VSS by
assuming a value for the COD/VSS ratio of the influent solids. See Melcer et al. (2003), Section 6.8 for
the calculation procedure.
METHOD 3: Bench-Scale Tests
Laboratory bench-scale analysis can provide accurate estimates of COD fractions as well as other
useful information such as TKN and TP fractions and nitrifier-specific growth rates. These analyses,
however, are labor intensive and time consuming (approximately 8 weeks). Melcer et al. (2003)
recommends that they be considered if plant data are not available or if results are unusual.
The recommended approach is to use a laboratory scale fill-and-draw sequencing batch reactor
(SBR) system. The designer should select a sludge age for the system that allows good solid liquid
separation (the example from Melcer et al. used a sludge age of 15 days). At startup, the SBR is seeded
with microorganisms from the activated sludge system. Wastewater is added to the system where it is
treated through the standard fill-react-settle-waste-draw SBR cycle. The SBR should be operated over
at least 2 to 3 sludge ages to achieve steady state. Intensive monitoring can then be used to
characterize COD and TKN fractions and nitrification kinetics. See Appendix D of Melcer et al. (2003) for
the full method description.
The cost of an SBR bench-scale test is estimated to be around $50,000 (Takacs 2009).
10.7.3 Data Checks
The following should be true for COD fractions for most municipal wastewaters:
• X: should be 20 to 40 percent of VSS.
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• The Ss (or rbCOD) fraction is higher when conditions in the collection system are favorable for
fermentation (i.e., warm climate, flat sewers, many force mains—see Chapter 5 for
background).
• When TSS in the plant influent is higher than normal, the particulate unbiodegradable COD
fraction may be higher than normal.
10.8 Characterization of Nutrient Fractions
10.8.1 Nitrogen
Total nitrogen in municipal wastewater is typically characterized by TKN. In some cases,
although rarely, nitrate or nitrite nitrogen (NOX) may be present, such as when recycle streams return
appreciable amounts of NOX to the head of the plant or the sewers have very steep slopes resulting in
rapid sheet flow within the sewers. NOX should then be included as an influent characteristic separate
from TKN.
Figure 10-6 shows the TKN fractions that are used in modeling. Table 10-3 provides a
description, the measurement method, and a range of expected concentrations for key nitrogen
fractions. Ammonia (both free and saline) makes up 60 to 75 percent of TKN, with the remaining
portion as organically bound biodegradable and unbiodegradable material. The ammonia/TKN ratio is
of particular importance in determining the effects of nitrification on pH. While nitrification of
ammonia to nitrate consumes alkalinity, hydrolysis of organically bound TKN adds alkalinity. If the
latter is not considered, the model could predict an unrealistically low pH.
Soluble unbiodegradable (inert) nitrogen, sometimes referred to as recalcitrant dissolved
organic nitrogen (rDON), has become increasingly important for plants trying to achieve low effluent TN
concentrations, as it passes through the plant and into the effluent. The higher the rDON, the lower the
concentrations of ammonia and NOx must be to achieve a certain effluent TN concentration.
WWTPs will usually have historic data for influent and effluent TKN or ammonia. Historical data
may only contain ammonia, which may be reported as total nitrogen, and additional tests may be
needed to determine the ammonia/TKN ratio. Plants may also have ammonia or nitrate concentrations
depending on permit requirements and monitoring practices.
Recommendations for estimating the remaining nitrogen fractions are provided below.
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Ammonia (SNH)
Soluble
Unbiodegradable
TKN (NOB)
Soluble
Biodegradable
TKN (SNB)
Particulate
Unbiodegradable
TKN (XNI)
Particulate
Biodegradable
TKN (XNB)
\
i
i
\
J
Soluble
Total TKN
r Particulate
J
Figure 10-6. TKN components for municipal wastewater.
Note: TKN fractions are not to scale and vary from plant to plant.
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Table 10-3. TKN Fractions in Municipal Wastewater
Fraction
Total Kjeldahl
Nitrogen
Ammonia (free and
saline)
Soluble
Unbiodegradable TKN
Soluble Biodegradable
TKN
Particulate
Unbiodegradable TKN
Particulate
Biodegradable TKN
Symbol1
TKNINF
SNH
NUB
(rDON)
SNB
XNI
XNB
Description
The total nitrogen load on
the plant
The total ammonia
Soluble Unbiodegradable
dissolved organic nitrogen
that passes through the
plant untouched
The portion of
biodegradable nitrogen that
is soluble
The portion of particulate
bound nitrogen that is not
biodegradable
The portion of particulate
bound nitrogen that is
biodegradable
Measurement
Method(s)
Directly measured using
colorimetric or titration
techniques
Directly measured
Difficult to determine,
default values often
used
Total soluble fraction
determined by filtering
sample, measuring TKN,
and subtracting
ammonia. The
biodegradable portion is
determined using
assumptions
Total particulate fraction
determined by filtering
sample and subtracting
the soluble TKN. The
Unbiodegradable
fraction is determined
using assumptions.
Total particulate fraction
determined by filtering
sample and subtracting
the soluble TKN. The
biodegradable fraction is
determined using
assumptions.
Typical
Concentration in
Municipal
Wastewater2 (mg/L)
25 -70 depending on
per capita water
consumption
20-30
0.5- 1.5. Higher values
when specific industrial
wastes are added
SNB + XNB = 0 - 10
2-8
SNB + XNB = 0 - 10
Notes:
1 The literature contains more than one symbol for some components. The symbols shown are generally consistent with
Melcer et al. (2003). Other commonly used symbols are included in parentheses. Note that the symbol shown represents
concentration, expressed as milligrams per liter (mg/L). Fractions of total COD are represented by the letter "F" and a subscript
to represent the influent COD fraction (e.g. Fs, is the fraction of the influent COD that is Unbiodegradable and soluble).
2 Derived from Melcer et al. (2003), Table 4-2.
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Ammonia (SNH)
When characterizing TKN, the most important fraction is the ammonia fraction, SNH. It is the
largest fraction and the component of TKN that is oxidized to nitrite then nitrate during nitrification.
Ammonia is measured directly using standard laboratory methods or with online analyzers.
Soluble Nitrogen Fractions
Soluble unbiodegradable nitrogen (SN!) is also often referred to as rDON. rDON is important
because it passes through the plant unreacted and can limit the amount of nitrogen removal that is
possible. Soluble biodegradable nitrogen (SNB) is organically bound nitrogen that has not yet been
converted to ammonia.
The total soluble nitrogen can be determined by subtracting the ammonia concentration from
the TKN concentration of a filtered sample. Unfortunately, there is no rapid method for experimentally
differentiating between SNB and rDON. The most straightforward way to determine rDON is to run a
pilot or bench scale test at a long SRT (> 12 days). Designers should check that all other requirements
for nitrification such as DO and pH are met and then determine the concentration of ammonia and TKN
in the effluent. The ammonia should be below 0.1 mg/L, and the bio-degradable organic nitrogen would
also be very low. The remainder of TKN is then rDON.
Site specific determination of rDON becomes very important for plants targeting low effluent
concentrations for TN (e.g., 3.0 mg/L). In other cases, default model values often suffice as the fractions
are generally a small percentage of the total (< 3 percent).
Particulate Nitrogen Fractions
Particulate nitrogen can be biodegradable (XNB) or unbiodegradable (XN!). As with the soluble
TKN fraction, it is not possible to directly measure the biodegradable and unbiodegradable fractions.
The particulate unbiodegradable portion, however, can be derived based on its relationship to
particulate unbiodegradable COD (X:) in the mixed liquor solids. The ratio of TKN to COD in the mixed
liquor solids is known to be approximately 0.07 mg N/mg COD. Therefore, XN! can be calculated as:
XN| = 0.07 Xj Eq. 10-4
As a check, results should be compared to total influent TKN. Melcer et al. (2003) report that XN! is
approximately 10 percent of influent TKN for municipal wastewater.
10.8.2 Phosphorus
Knowing the phosphorus fractionation is important for plants that provide biological
phosphorus removal. The P/VSS ratio is important in sizing the anaerobic zone and will also be a factor
in designing the denitrification process. Figure 10-7 shows the fractions of total phosphorus. Table 10-
4 provides a description, the measurement method, and a range of expected concentrations for key
phosphorus fractions. All fractions are expressed as phosphorus in mg P/L
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Orthophosphate
(Spew)
Soluble
Unbiodegradable
Phosphorus (SP|)
Soluble
Biodegradable
Phosphorus(SPB)
Particulate
Unbiodegradable
Phosphorus (XP|)
Particulate
Biodegradable
Phosphorus (XPB)
Soluble
Total P
r Particulate
J
Figure 10-7. Phosphorus components in municipal wastewater.
Note: TP fractions are not to scale and vary from plant to plant.
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Table 10-4. Total Phosphorus Fractions in Municipal Wastewater
Fraction
Total Phosphorus
Orthophosphate
Soluble
Unbiodegradable
Phosphorus
Soluble Biodegradable
Phosphorus
Particulate
Unbiodegradable
Phosphorus
Particulate
Biodegradable
Phosphorus
Symbol1
TP
SpO4
SPI
SPB
Xpi
XPB
Description
The total phosphorus load
on the plant
The total orthophosphate
Soluble Unbiodegradable
phosphorus
The portion of
biodegradable phosphorus
that is soluble
The portion of particulate
bound phosphorus that is
not biodegradable
The portion of particulate
bound phosphorus that is
biodegradable
Measurement
Method(s)
Measured as ortho-
phosphorus following
chemical conversion
Directly measured
Total soluble fraction
measured for filtered
sample, estimate for
Unbiodegradable portion
Total soluble fraction
measured for filtered
sample, estimate for
biodegradable portion
Total particulate fraction
measured by subtracting
soluble from total,
estimate for
biodegradable portion
Total soluble fraction
measured by subtracting
soluble from total,
estimate for
biodegradable portion
Typical
Concentration in
Municipal
Wastewater2 (mg/L)
4-15
2-12
With XpB 0-10
1- 4
With SpB 0 - 10
Notes:
1. The literature contains more than one symbol for some components. The symbols shown are generally consistent with
Melcer et al. (2003). Other commonly used symbols are included in parentheses. Note that the symbol shown represents
concentration, expressed as milligrams per liter (mg/L). Fractions of total COD are represented by the letter "F" and a subscript
to represent the influent COD fraction (e.g. Fs, is the fraction of the influent COD that is Unbiodegradable and soluble).
2. Derived from Melcer et al. (2003), Table 4-2.
Orthophosphate
Orthophosphate is determined by direct measurement. Along with soluble biodegradable
phosphorus, it represents the phosphorus that will be taken up by PAOs in the aerobic zone. Complex
soluble phosphorus and some particulate phosphorus are hydrolyzed to mostly orthophosphate in the
first stage of treatment. Orthophosphate typically makes up a large portion of TP in municipal
wastewater (50 to as high as 80 percent). After primary sedimentation, nearly all of the remaining
phosphorus is rapidly converted to orthophosphorus in the first bioreactor. Almost all of the effluent
soluble phosphorus is orthophosphate.
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Soluble Fractions
Soluble fractions of phosphorus represent the organic phosphorus that is not particulate bound.
The biodegradable portion (SPB,) represents organically bound phosphorus that is taken up into the
biomass for growth. The unbiodegradable portion (SPi) is finely colloidal or soluble organic compounds,
some of which are associated with the soluble unbiodegradable COD. The total soluble portion can be
determined by filtering a sample and measuring total phosphorus in the filtrate. Sp:and SPB are not
easily differentiated. Generally a default value can be used for the fraction of TP that is SPL SPB is then
determined by subtracting SPL
Particulate Fractions
The particulate fractions are organically bound phosphorus within particles. The biodegradable
portion (XPB) is associated with the biodegradable particulate COD. The unbiodegradable portion (XP!) is
associated with unbiodegradable COD. The total particulate fraction is determined by subtracting the
total soluble fraction as determined above from the total phosphorus. XP! can be determined as it is a
given fraction of the unbiodegradable COD. Typically a value of 0.02 to 0.03 is used. XPB can then be
determined by subtracting XP!from the total particulate phosphorus.
10.9 Kinetic and Stoichiometric Parameters
Mathematical models require estimates of kinetic and Stoichiometric parameters such as yield
and growth rates to predict performance of wastewater treatment processes. Model default values for
these parameters are based on thousands of observations and should generally be acceptable for most
plants. The exception is nitrification rates.
The maximum specific nitrifier growth rate, U.AUT, and the endogenous nitrifier decay rate, bAUT,
are extremely important in the design of activated sludge nitrification systems. This is because the net
growth rate (u.AuT-bALn-) determines minimum SRT needed for complete nitrification to occur during
steady state conditions. In equation format,
SRT = l/(HAUT-bAUT) Eq. 10-5
Where:
SRT = Solids residence time (days)
H-AUT, = Specific nitrifier growth rate (d"1)
bAUT = Endogenous nitrifier decay rate (d"1)
(Source: adapted from WEF and ASCE 2006, equation 3.8)
The literature reports little variation in endogenous nitrifier decay rate, bAUT, from plant to
plant. Previously, bAUT was considered to be low compared to u.AUT and, for this reason, it was excluded
from much of the previous literature on nitrification rates. Recent results summarized by Melcer et al.
(2003) show that experimental methods underestimated bAUT and that it is more significant than
previously reported (bAUT was previously thought to be between 0 and 0.05 d"1; however, new data
showed that it could be as high as 0.2 d-1at 20°C). Underestimates of bAUT have resulted in
underestimates of u.AUT by a factor of 2 in some cases (Melcer et al. 2003). Although not an issue for
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completely nitrifying systems, an underestimate in u.AUT could lead to conservative predictions of
performance for low SRT systems. See Chapter 4 for additional discussion of nitrification kinetics.
Melcer et al. includes new methods for site specific determination of bAUT. In the absence of
more specific data, the authors provide the following as a reasonable estimate of bAUT:
bAUT= 0.17 * 1.029(T"20) d"1 Eq. 10-6
Where:
T = Temperature in °C
d = Days
Although bAUT is believed to be fairly consistent from plant to plant, researchers have observed
significant variation in u.AUT, with measured values at 20 °C (u.AU2o) of 0.2 to 1.0 per day. Factors affecting
the rate of nitrification include industrial input, dissolved oxygen concentration, temperature, alkalinity,
pH, and hydraulics within the reactor. Recent studies report u.AU2o ranges from 0.7 to 1.0 if no inhibition
is experienced (Melcer et al. 2003). See Chapter 4 for an in-depth discussion of the nitrification process.
Because of its importance in designing nitrification systems, u.AUT should be based on site-
specific characterization of the influent wastewater. Melcer et al. (2003) presents three rapid methods
for measuring u.AUT: Low F/M SBR, High F/M, and the washout method. Each method has its distinct
advantages and disadvantages. The methods have produced the same results when conducted in
parallel; therefore, the choice of the method depends on the preferences of the designer. It is
important to note that each of these three methods determines the net growth rate (u.AUT - bAUT), not
u.AUT explicitly. bAUT must therefore be estimated separately or assumed.
In some cases, it may be possible to use the model to estimate u.AUT by adjusting it until model
predictions match observed effluent quality (e.g., ammonia concentration) from the nitrification
process. This iterative approach requires extensive data collection and is only meaningful when there
are diurnal ammonia influent and effluent data. In fully nitrifying plants operating at high SRTs, the
effluent ammonia will always be low, and this method would not apply.
10.10 Calibration
Calibration is the process of adjusting simulation parameters to match performance of a WWTP.
Calibration has three main objectives, as summarized below.
1) To minimize the error between predicted and measured performance: The acceptable level of
agreement should match objectives (see Section 10.4.2). Typical goals are no more than 10 to
15 percent difference for initial design and 5 percent for final design, although larger deviations
during dynamic simulations are commonly accepted. Error can be determined either by visual
analysis of observed and predicted data or statistical analysis. A perfect match is not achievable
and should not be a goal of calibration.
2) To define the accuracy of the model: Based on the final agreement between measured and
predicted values, the expected accuracy of the model can be defined for individual unit
processes and constituents (e.g., COD, N).
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3) To establish the circumstances under which the model can be used to simulate the treatment
plant, also referred to as the model "design space": A model should not be used for conditions
for which it has not been calibrated to actual plant performance (e.g., extreme flow conditions
such as a storm event).
The calibration process depends on the extent and quality of available data and the complexity
of the existing treatment processes. The intended use and goals of modeling define the level of
calibration (i.e., the acceptable error and whether or not the model needs to be calibrated to dynamic
operating conditions). Good calibration requires detailed understanding of the fundamental treatment
processes at the plant and of the principles of the model. Overriding principles of calibration are
provided in the text box on the next page.
In an ideal situation, historical data for influent loading and plant operations are accurate and
meet calibration needs. However, this is rarely the case. As noted earlier in this chapter, a tiered
approach is recommended whereby designers first attempt to calibrate the model with existing data
then collect additional data based on calibration results. Data checking and additional data gathering
occur throughout the calibration procedure. This iterative approach is often time consuming and will
require several rounds of data checking, operator interviews, and in some cases additional sampling.
Several references (e.g., Petersen et al. 2002; Hulsbeek et al. 2002) report that the time required for
calibration is often longer than expected.
Calibration should follow a stepwise, iterative procedure starting with sludge production and
proceeding to composition, aeration, nitrification, denitrification, and biological phosphorus removal.
Mass balances are an integral part of calibration and are incorporated throughout the procedure.
Designers should first calibrate the model for steady state conditions (e.g., average concentrations in
influent, effluent, and waste streams during normal dry weather conditions), then calibrate it for
dynamic conditions. For dynamic simulations, it is important to initialize the model properly to establish
operating conditions consistent with observed data to be used for calibration.
An overall recommended approach to calibration, including mass balance checks, is provided
below. The recommended approach has 6 steps, assuming that the reader has already gone through
the data checks in Chapter 7. The approach is based on a review of several publications on the topic
(Third et al. 2007; Gernaey et al. 2002; Melcer et al. 2003; Hulsbeek et al. 2002; and Petersen et al.
2002). The 6-step calibration approach is not the only valid one—alternative approaches can work just
as well as long as designers adhere to the overriding principles of calibration as stated in the text box on
the following page. Readers should also consult the GMP Unified Protocol when it becomes available
from IWA. Technical papers such as McConnell et al. (2008) and Latimer et al. (2008) provide additional
examples calibration experiences and may be useful references for the reader.
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Principles of Calibration
• Use a process engineering approach. Adjust parameters one by one based on the experience
of the designer and an in-depth understanding of fundamental wastewater treatment
processes.
• If the modeled treatment plant performance deviates substantially from observed values, it
may result from errors in the hydraulic model (e.g., flow splits, recycle flow rate, wastage rate
not properly configured). The problem could also be major operational set points such as
minimum DO concentration and aeration rate. Operational data have a large influence on
model outputs compared to wastewater characterization and kinetic/stoichiometric
parameters (Gernaey et al. 2003). Another problem may be changes in the influent
characteristics that are not modeled such as chlorine addition in the collection system for odor
control.
• Adjust kinetic and stoichiometric parameters only as a last resort. The default kinetic and
stoichiometric values in models represent averages of thousands of measured values from
hundreds of plants. With few exceptions (see below), they should not deviate substantially
from plant to plant. If predicted results do not agree with measured values, designers should
always first ask the questions, "Is there something happening at the plant that could cause this
result?" and "What are the possible reasons that the plant is not performing in the model as
predicted?" The exception is the maximum specific nitrifier growth rate (U.AUT), which could be
very low if there is nitrifying inhibitors in the waste stream or higher if fewer than average
inhibitors are present.
• Justify and document changes to kinetic and stoichiometric parameters and estimated
constituent fractions.
Step 1. Check Process Configuration and Plant Hydraulics
• Compare design capacities of each unit process to modeled outputs with maximum operating
capacities (Third et al. 2007).
• Perform total flow balance for the WWTP. Most simulators will automatically perform a flow
balance around the plant. In simple terms, influent flow should equal effluent flow plus waste
stream flows.
• Perform flow balances for each unit process considering wastage and recycle flows.
Step 2. Calibrate Sludge Production and Characteristics
• If not done already, perform mass balance for COD and TSS/VSS around the primary clarifier
(see Section 10.7.2, Method 1).
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• Perform mass balance of the inert fractions through the plant. Because it is not changed during
biological reactions, the mass of inert solids in and out of the plant should balance. Take into
account that phosphorus and magnesium removed either by chemicals or biology will register
as inert material. Also, the metals used for precipitation would add to this fraction. If the inert
mass of the solids in the waste products of the existing plant does not balance, check the
following:
- Wastage rate (this is a very common problem)
- RAS flow rate
• The predicted mass balance for total solids in the simulator model should correlate with that in
the full-scale plant, taking into account the issue of inert solids as discussed above.
• A mass balance for nitrogen is not possible because of the denitrified nitrogen that is released
to the atmosphere.
• Perform mass balance for phosphorus, which cannot be destroyed and has no gaseous phase,
by measuring influent, effluent, and wasted sludge concentrations. Results can be used to
check the mass of sludge wasted.
Step 3. Calibrate the Aeration Process
• A COD balance can be used to check the oxygen uptake at the WWTP. For the following
equation, solve for OUR:
CODT,INF = CODe + OUR + (Nd * 2.86) + Qs * Gs,org * 1.42) - (4.56 * Nn) Eq. 10-7
Where:
CODT,INF = COD load in the influent (kg O2/d)
CODe = COD load in the effluent (kg O2/d)
OUR = Oxygen uptake rate (kg O2/d)
Nd = Total denitrified nitrogen load (kg/d)
2.86 = Oxygen reduction-equivalent (1 kg NO3-N is equivalent to 2.68 kg O2)
Qs = Daily volume of excess sludge (m3/d)
Gs,org = Concentration of VSS in the excess sludge (kg VSS/m3)
1.42 = Conversion faction from VSS to COD equivalents (kg COD/kg VSS)
4.56 = Conversion factor for calculation of oxygen uptake during nitrification
Nn = Total nitrified nitrogen load (kg/d)
(Hulsbeek et al. 2002, equation 3)
Designers can use the estimated OUR to check the aeration efficiency of the plant (kg O2 / kWh)
or to detect errors in the setup of aerators.
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Step 4. Calibrate the Nitrification Process
• Compare observed effluent NH3-N concentrations to simulation results. If results do not match,
check the following:
Predicted vs. observed DO concentration in aeration basin
Basin hydraulics for short circuiting
- Alkalinity depletion
Method for determining maximum nitrifier specific growth rate and decay rate (see section
10.9)
Recheck sludge production (Step 2).
Step 5. Calibrate the Denitrification Process
• Make sure that the model function for carrying DO from one basin to the next is on. This will
model the transfer of oxygen in the recycled mixed liquor.
• Compare observed effluent NO3-N concentrations to simulation results. If results do not match,
check the following:
Denitrification in return sludge
Re-aeration of denitrification process (e.g., from pumps or overflow weirs)
If this does not solve the problem, consider modifying the anoxic reduction factor or possibly
the heterotrophic decay factor (Hulsbeek et al. 2002).
Recheck sludge production (Step 2).
Step 6. Calibrate the Biological Phosphorus Removal Process
• Compare observed effluent TP concentrations to simulation results. If results do not match,
check for the following:
DO and nitrate in the anaerobic zone
rbCOD in the anaerobic zone
- Addition of oxidizing agent such as chlorine to the influent that will result in inhibition of
fermentation in the anaerobic zone
10.11 Validation
Whereas calibration ensures the model fits the original data set, validation seeks to ensure that
the model is valid over a wider range of conditions. Validation includes running the model with a second
set of data, different than the original calibration data. It also includes checking the model results to
ensure they make sense with the physical system. The new set of data does not need to be as extensive
as the original calibration data. An additional set of typically measured wastewater influent
characteristics and flows along with effluent data should be sufficient. The data should be under
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different conditions than the calibration data, but should still remain within the conditions under which
the model is valid.
The new data should be entered into the model and the model run. If the model predicts the
effluent data within an acceptable error, the model can be considered validated. If not, the model
calibration procedure may need to be repeated. The following are common checks that should be
conducted with the new data set to validate the model.
• Check mass balances. Using the new set of data, check the mass balances for flow and
solids again. If an error is found, the sludge wastage rate and sludge composition are factors
that should be examined.
• Check nutrient balances. Check the balances for phosphorus and nitrogen again with the
new data. Model parameters to check include the nitrifier growth rate, the nitrogen
content of inert particles, and the ammonia and oxygen half saturation constants for
heterotrophs and autotrophs.
• Check alkalinity balance. The difference between the influent alkalinity and the effluent
alkalinity should equal the net consumption of alkalinity from the biological reactions.
• Check solids production and net solids yield. If these values are off, they could indicate
errors in the sludge wastage rate or solids content.
• Check nitrification and denitrification rates. Reasonable values of rate constants can be
found in Jeppsson (1996) and Hulsbeek et al. (2002).
• Check residence times (SRT and HRT). The residence times in the reactors should be
checked using the validated data. If the times are not reasonable and do not correlate with
actual plant data, the model should be recalibrated.
10.12 Simulation of Design Alternatives for Nutrient Removal
Once the designer has calibrated and validated the model, the next step is to set up and run
design scenarios for biological nutrient removal that answer the questions laid out in the modeling
objectives. In simple terms, the influent flow and wastewater characteristics are modified to match
design conditions, and the configuration of the existing plant is modified to incorporate candidate
treatment processes as identified in Chapter 8 of this design manual. Designers should then conduct
the following simulations to determine if the candidate treatment processes meet the design objectives
(Rieger2009):
• Steady state simulations, checking for:
Overall mass balancing
Long-term performance by evaluating effluent quality
• Dynamic simulations as follows:
Simulate diurnal peaks to determine equipment limits
Determine set points for controller design
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Simulate seasonal changes to establish long-term operational strategies
Simulate storm events to determine optimal handling
It is in this step that designers realize the benefits of the effort required to set up and calibrate a
detailed simulator of the existing WWTP. A large number of operating adjustments and treatment
alternatives can be rapidly simulated to determine their effectiveness. Dynamic modeling is an
extremely powerful tool in determining the most efficient operations for biological nutrient removal.
For the most promising biological nutrient removal technologies evaluated, designers should
summarize information on long-term performance during a variety of operating conditions, if possible,
and operating requirements (e.g., energy requirements, chemical addition, etc.).
10.13 Additional Procedures for Design
This section provides recommendations for design, equipment specification, and operation of the
following nutrient removal technologies:
• Sequencing Batch Reactors
• Denitrification Filters
• Primary Sludge Fermenters
10.13.1 Sequencing Batch Reactors (SBRs)
SBRs are fill and draw systems that combine equalization, biological treatment, and secondary
clarification in one tank using a time control sequence. As noted in Chapter 6, they can be designed and
operated for nitrogen and phosphorus removal by cycling through aerobic, anoxic, and anaerobic
conditions (USEPA 1999). Recommendations for the design and operation for nutrient removal are
presented below.
• Designs should have a minimum of two SBR basins to allow for maintenance, handling of high
flows, and seasonal variations. Designs should also consider influent equalization basins with
capacity to hold peak flows during one treatment cycle.
• For nitrogen removal, an initial anoxic phase (mixing only, no aeration) should be followed by an
aerobic phase, mimicking the Modified Ludzack Ettinger (MLE) process. Multiple anoxic phases
within a single SBR cycle can improve performance (i.e., a second anoxic phase after the aerobic
phase can be used to mimic a 4-stage Bardenpho process).
• Alkalinity may need to be added to keep it in the range of 40 to 70 mg/L as CaCO3. Methanol or
another external organic carbon source may need to be added for denitrification to maintain a
relatively high rate of denitrification, although this is not common.
• SBRs can be designed for biological phosphorus removal (BPR) by using a large batch reactor
with a reseeding anaerobic zone (WEF and ASCE 2006). If the biodegradable COD to TKN plus TP
ratio is sufficiently large, BPR can be accomplished by maintaining non-aerated conditions after
all of the nitrates have been denitrified, i.e. after the system has become truly anaerobic. It is
necessary for VFAs to be available during the anaerobic period for BPR to occur. The effluent
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soluble phosphorus concentration will be determined by the VFA to TP ratio during that period
of time, assuming that conditions are suitable for phosphorus uptake during the subsequent
aerated period. Many SBR systems successfully remove phosphorus by adding a metal salt
either upstream of the SBR, during the react phase, or following the SBR if effluent filtration is in
place (Young, 2008).
• Typical design parameters are as follows (WEF and ASCE 2006, Table 5.2):
SRT (including anaerobic and anoxic zones): 20 to 40 days, although SRTs considerably less
than 20 days can be used. The primary purpose of long SRTs is to minimize the production
of WAS
- MLSS: 3to4g/L
HRT (in terms of average flow):
• Anaerobic period: 1.5 to 3 hrs
• Anoxic period: 1 to 3 hrs
• Aerobic period: 2 to 4 hrs
• Settling period: 0.5 to 1 hr
• Decant period: 0.2 to 1 hr
Note that an anaerobic period cannot be obtained in an SBR where nitrification is obtained until
all of the nitrates have been reduced to nitrogen gas. Influent BOD is used for nitirifcation
resulting in reduction of rbCOD needed for BPR.
Some simulator packages have models for simulating a variety of SBR processes. The input of
wastewater data and plant specific characteristics such as nitrification rates can be introduced as for
modeling the flow through process. Because an SBR is never at steady state, diurnal modeling should be
used. The simulator may need to be run for a long period to ensure stabilization.
10.13.2 Denitrification Filters
Traditional design of denitrifying filters is done by first calculating a required surface area based
on an assumed loading rate. Loading rates of 1.5 to 2 gpm/ft2 and media depths of 6 to 9 feet are
commonly used. For down-flow filters, the backwash is designed to minimize the amount of air
introduced. The influent weir is also designed to avoid introduction of air. Upflow filters have the
advantage of not requiring backwash. The methanol feed system should be designed to adjust the feed
of methanol based on the COD in the influent. EPA has prepared a fact sheet on denitrification filters
that includes the following information (USEPA 2007):
• Filter design characteristics, including:
Influent weirs
Media
Underdrain system
Nitrogen release cycle
Backwashing and filter controls
Methanol feed control system
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• A comparison of denitrification filter manufacturers and equipment
• Cost factors
The fact sheet can be accessed at http://www.epa.gOV/OWM//mtb/mtbfact.htm .
10.13.3 Primary Sludge Fermenters
Section 5.3.2 of this design manual describes the wide variety of approaches that can be used to
achieve fermentation of primary sludge for BPR. The following should be considered as general
guidelines for design and equipment specification regardless of configuration (WEF and ASCE 2006;
Chanono 2006; Chen 2004):
• Sludge collector designs should be higher torque than conventional designs due to increased
solids loading.
• The maximum SRT should depend on the temperature to avoid methanogenesis. It may range
from 2 to 8 days.
• Air spargers may be used to reduce methanogenesis.
• A temperature between 16 and 25 degrees Celsius has been found optimal but many fermenters
work at lower temperatures. Covering and insulation in cold climates has merit.
• The sludge concentration into the thickener will depend on the flow rate of the primary sludge.
A continuous flow of around 2 to 4 percent of the influent is recommended to allow liquid for
VFA elutriation. In a static fermenter the sludge could be thickened to 6 percent solids provided
the sludge pumps can handle the dense sludge. The denser the sludge the more VFA will be
elutriated.
• Because of the potential to produce carbon dioxide and hydrogen sulfide, corrosion-resistant
materials should be used.
• Primary and fermented sludge pumps should generally be of the positive displacement type.
Recessed impeller centrifugal pumps can be used for lower solids content streams such as
fermentate.
• Sludge grinders or screens should be installed on sludge recycle lines to prevent clogging of the
pumps.
• For complete mix tanks, slow speed mixers that impart 8 to 10 W/m3 energy into the liquid are
recommended.
• Scum removal should be provided on mixing tanks and thickeners.
• Thickeners should be covered and the headspace scrubbed to control odor.
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• Measuring oxidation reduction potential and keeping it above -600 mV by intermittently
bubbling air through may help prevent methanogenesis in completely mixed fermenters.
Adding some nitrified effluent to the static fermenter feed will serve the same purpose
• Installation of hydrogen sulfide sensors and alarms is a good idea to ensure worker safety and
prevent methanogenesis since both tend to be generated at the same oxidation reduction
potential.
• Sludge density meters and pH meters can aid in process control. A pH level of around 5 is an
indication of good acid fermentation.
10.14 Design Checks for Biological Nitrogen and Phosphorus Removal
Tables 10-5 and 10-6 provide key design checks for biological nitrogen and phosphorus removal
systems, respectively. These tables synthesize information presented throughout this design manual
and can be used at various steps in the design process to ensure that key factors are considered.
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Table 10-5. Design Checks for Biological Nitrogen Removal
Question
Yes
No
1. Is the nitrification process designed for cold weather conditions?
The rate of nitrification is significantly impacted by temperature. It will decrease
by approximately half for every 8 to 10°C decrease in operating temperature.
2. Does the nitrification system design consider the impact of diurnal loadings of
ammonia?
Influent ammonia concentrations vary, with peaks occurring during the day and
much lower concentrations occurring overnight. Peak loads during the day are
typically 1.5 to 2.0 times the average with some observations as high as 2.8
times the average.
3. Does the nitrification system design account for recycles, including realistic
predicted constituent loadings, flow rate, and schedule?
Return streams from sludge dewatering operations can add significant ammonia
to the head of the plant. In many plants, supernatant from sludge dewatering is
discharged during the day, which coincides with peak ammonia loadings to the
plant.
4. How does the design of the nitrification system take into account wet weather
flows?
Wet weather flows can significantly increase hydraulic loading to the treatment
plant, change the nature of the influent wastewater, and potentially washout
the nitrification system. Strategies for managing wet weather flows range from
collection system improvements to in-plant designs such as equalization basins.
5. Is the operating DO for nitrification at least 2.0 mg/L?
In general, the DO at the influent end of the aeration basin should be at least 2.0
mg/L. For longer SRT systems, it can taper to 0.5 mg/L at the end of the basin.
There are many benefits to designing flexible aeration systems that allow
operators to reduce the DO at the end of the basin after the ammonia has been
oxidized and less oxygen transfer is required. Integrated Fixed-film Activated
Sludge (IFAS) systems may require higher DO concentrations in the mixed liquor
to maintain the oxygen needed for attached biomass if maximum nitrification
rates are desired. However, if simultaneous denitrification is desired in an IFAS
system, lower DO concentrations can both stimulate denitrification and reduce
the oxygen requirement in the aerated bioreactor because of COD stabilized
using NOxas the electron acceptor. This can be an effective method of nitrogen
removal and can remove as much as 30% of the influent nitrogen. In long
nominal hydraulic retention time systems such as oxidation ditches or extended
aeration systems, DO concentrations between 0.25 and 0.50 mg/L can produce
excellent simultaneous nitrification/denitrification in the same bioreactor
because the rate of oxygen transfer does not control the rate of nitrification in
such a process, stoichiometry does( i.e. the balance between electron donors
(BOD) and electron acceptors (DO, NOX) in the bioreactor).
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Table 10-5. Design Checks for Biological Nitrogen Removal (Continued)
Question
6. Does the design consider impacts of nitrification on alkalinity and pH?
The nitrification process consumes alkalinity. If the influent alkalinity is low,
nitrification may reduce the mixed liquor pH, which can have an impact on
performance because pH values below 7.0 significantly inhibit nitrification and
pH values below 6.5 strongly inhibit nitrification and may result In nitrifier
washout. Denitrification, however, partially replenishes alkalinity (50 percent of
the alkalinity destroyed by complete nitrification of 1 mg/L ofNH4-N is restored
by denitrification of 1 mg/L NO3-N). Addition of lime, sodium hydroxide, or soda
ash can be used to add alkalinity if low alkalinity is determined to be a problem.
7. Is the design of the nitrification system based on site-specific determination of
the nitrification rate?
The nitrification rate is extremely important because it determines the minimum
SRT needed for complete nitrification to occur during steady state conditions.
The nitrification rate varies significantly from plant to plant and can be low if
inhibitors such as heavy metals are in present. Melcer et al. (2003) presents
three rapid methods for measuring maximum specific nitrifier growth rate, |o,AUT.
8. Does the design minimize the amount of oxygen in the denitrification system?
Design and operation methods that minimize DO quantities entering the anoxic
zones can reduce the amount of exogenous carbon needed and provide a more
stable operation.
9. If an external carbon source is added for denitrification, does the design
include an analysis of possible increased sludge volume?
If an external carbon source is added to improve denitrification, there may be a
small increase in waste sludge production. The magnitude of the increase
depends on the amount and type of organic carbon added. Nitrification
processes usually decrease the amount of sludge produced, however, so net
production for the design of a new combined nitrification/ denitrification system
may be equal to or less than sludge production by the same size conventional
activated sludge plant.
10. If the design calls for methanol to be added for denitrification, are proper
safety and handling procedures considered?
Methanol is corrosive and combustible and requires special handling to meet
Occupational Safety and Health Administration (OSHA) and other requirements.
11. If biological or chemical phosphorus removal is also being practiced, does the
design consider the possible impacts of phosphorus depletion on nitrification
and denitrification?
Phosphorus is a necessary nutrient for microbial growth. IfTP is reduced too low
before nitrification or denitrification, the growth of the microorganisms could be
inhibited, thereby requiring the addition of supplemental phosphorus. This is
most likely to occur when tertiary denitrification filters are used, but inhibition of
nitrification also has been observed in plants practicing chemical phosphorus
removal in the primary clarifiers. It is also important to note that both BPR and
denitrification require BOD to drive the reactions. If BPR is consuming some of
the BOD, there will be less for denitrification.
Yes
No
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Table 10-6. Design Checks for Biological Phosphorus Removal
Question
Yes
No
1. Is the rbCOD:TP ratio of the influent consistently 15:1 or higher?
If not, consider a strategy to increase rbCOD such as fermenting primary sludge or
adding a commercial product (e.g., acetic acid).
2. Are odor control strategies used in the collection system (e.g., chlorination,
addition of nitrate salts)?
If yes, the collection system managers should use strategies that do not prevent
fermentation ofrbCOD to VFA (e.g., ferric iron addition for odor control or
modifying the pH to between 9 and 10).
3. In situations where BPR is operated without nitrification, is the design SRT for
the entire BPR system at least 3 to 4 days in temperate regions?
The minimum SRT to prevent wash-out of BPR is a function of the mixed liquor
temperature, and BPR can be accomplished at SRTs less than 4 days if the
temperature is high. Because BPR is driven primarily by microbial population
dynamics resulting from the available VFA to available TP ratio in the anaerobic
zone, SRT values above the minimum value up to 25 days or more have little effect
upon the phosphorus removal accomplished by the process. Population dynamics
also result in better steady-state performance of BPR at low temperatures
because the primary competitors of the PAOs, the glycogen accumulating
organisms (GAOs), are more strongly inhibited by temperature than PAOs.
4. Does the design SRT for fermenters consider operating temperature?
VFA formation in the sewers is temperature dependent. Check the winter
production. Higher SRTs (> 4 days) may be needed in colder climates to account
for slower fermentation of rbCOD to VFA.
5. Does the design take into account wet weather flows?
As with all biological wastewater treatment processes, BPR works best under
steady state conditions. Equalization basins can be very effective at managing
storm flows as well as nutrient rich recycle streams. Designs should also consider
flexibility in aeration basins to take basins or compartments on and offline in
response to seasonally changing flows.
6. Does the design allow for flexibility of operations?
Most plants initially operate well below the design load. The design should allow
for the flexible operation of various anaerobic, anoxic, and aerobic zones to match
changes in operating conditions (e.g., turn down on blowers, recycle pumps, or
less mixing energy in the aeration basin). A good approach to use when flows and
loads are low is multiple treatment trains, of which one or more can be taken out
of service.
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Table 10-6. Design Checks for Biological Phosphorus Removal (Continued)
Question
Yes
No
7. Does the design HRT of the anaerobic zone consider secondary release?
// the HRT of the anaerobic basin is too long, secondary release of phosphorus can
occur after depletion of VFA. The anaerobic HRT is recommended to be 5 to 15
percent of the total HRT of the BPR system. HRTs on the lower side may be used in
warm climates and on the higher side in colder climates. Note, however, that the
development of anaerobic conditions in the latter part of the anoxic zone typically
does not cause secondary release, but adds to the VFA uptake accomplished in the
anaerobic zone, resulting in greater phosphorus removal by the total system. Also
note that an oversized aerobic zone will result in excessive glycogen depletion in
the PAOs, resulting in less VFA storage in the anaerobic zone and less overall
phosphorus removal.
8. Does the design take into account DO or nitrates in recycle streams to the
anaerobic zone?
The design should minimize introduction of dissolved oxygen or nitrates to the
anaerobic zone from recycle streams. Any DO or NOx in the anaerobic zone will
reduce VFA storage by the PAOs, and, thus, the total phosphorus removal
obtained.
9. Are measures in place to prevent backmixing?
If the anaerobic zone is followed immediately by an anoxic or aerobic zone, the
configuration should prevent mixed liquor from seeping back into the anaerobic
zone. Aerated zones have a higher water surface that unaerated zones. Use an
overflow baffle between zones or an underflow baffle with openings small enough
to maintain a forward velocity of at least 1 foot per second. Modeling can also be
used to ensure that the configuration, mixing rate, and overall hydraulics are such
that no backmixing occurs.
10. Will chemical addition or nitrification cause the pH of the anaerobic zone to
regularly fall below 6.9?
If yes, consider adding alkalinity to increase pH to more than 7.0. However, note
that BPR can be accomplished at pH values in the anaerobic zone as low as 6.3,
but at significantly slower rates than at pH values about 7. If the anaerobic zone
pH is less than 6.0, BPR will be completely inhibited. A common pH profile for a
BPR process is 6.8 to 7.0 in the anaerobic zone, but 7.8 to 8.2 in the aerobic zone,
particularly if the system includes an anoxic zone between them.
11. Does the design take into account phosphates in return streams from sludge
dewatering?
In BPR plants, anaerobic digestion of sludge can release phosphorus into the liquid
stream, which is normally returned to the plant after sludge dewatering. Much of
the released phosphorus can, however, precipitate within the digester as struvite,
calcium phosphate precipitates, or vivianite. Dissolved air flotation (DAF) or
mechanical thickening can be used instead of gravity thickening to reduce
phosphorus release during WAS processing.
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Table 10-6. Design Checks for Biological Phosphorus Removal (Continued)
Question
12. If phosphorus removal is accompanied by nitrification in a single sludge system,
does the design factor consider nitrate in the RAS and its impact on BPR?
Nitrate can act as an electron acceptor for some heterotrophs, allowing them to
use VFA in the anaerobic zone in the absence of oxygen. It will also prevent
fermentation of the rbCOD to VFA. The model will generally account for this.
13 Does the design take into account possible struvite formation?
When BPR is practiced, struvite may precipitate in anaerobic sludge handling
processes. Depending on where they form, struvite crystals can plug centrifuge
ports and the pipes and pumps used to convey sludge. Some models predict
struvite formation, but not the impacts of it. Note that models do not capture all
phosphorus precipitation or adsorption mechanism in the digester and thus may
under-predict struvite precipitation and over-predict phosphate in the recycle
stream.
Yes
No
10.15 References
Albertson, O.E. 1995. Is CBOD5 Test Viable for Raw and Settled Wastewater? Journal of Environmental
Engineering. 121(7): 515-520.
Alex, J., M. Wichern, V. Spering, N. Halft, M. Ahnert, T. Frehmann, I. Hobus, G. Langergraber, M. Plattes,
S. Winkler, and D. Woerner. 2007. A Method to Use Dynamic Simulation in Compliance to Design
Rules to Refine WWTP Planning. In Proceedings 10th IWA Specialised Conference on Design,
Operation, and Economics of Large Wastewater Treatment Plants. September 9-13. IWA.
American Public Health Association (APHA), American Water Works Association (AWWA) & Water
Environment Federation (WEF). 2005. Standard Methods for the Examination of Water and
Wastewater, 21st Edition. Eds. A.D. Eaton, LS. Clesceri, E.W. Rice, A.E. Greenberg, and M.H. Franson.
Aquifas. 2009. Aquifas - Unified Model for Activated Sludge, Membrane Bioreactor - MBR, IFAS, MBBR
and BAF. Cupertino, CA: MEC Technologies. Available online: www.aquifas.com
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11. Design Approach for Effluent Filtration
Chapter 11 covers:
11.1 Introduction
11.2 Selection of Filtration Technology
11.3 Granular Media Filters
11.4 Cloth Media Filters
11.5 Low-Pressure Membranes
11.6 Emerging Filtration Technologies for Phosphorus
Removal
11.7 References
11.1 Introduction
Effluent (or tertiary) filtration is a process to physically separate solids from wastewater during
tertiary treatment, either with or without tertiary clarification. Tertiary filtration is used for effluent
polishing and can be very effective in removing particulate and colloidal phosphorus from wastewater.
When combined with chemical treatment, tertiary filtration can achieve very low total phosphorus (TP)
levels in plant effluents. Tertiary filtration also reduces total suspended solids (TSS) to very low levels,
which in turn can result in more efficient disinfection and allow for the possible use of alternative
disinfectants to chlorine such as ultraviolet light (UV) (EPA Region 10 2007). Turbidity is often used as
the measure of tertiary filter performance instead of TSS because turbidity can be monitored
continuously using online instruments.
Filtration technologies for wastewater treatment are continuing to evolve. Historically, granular
media filtration was the most common type used at wastewater treatment plants (WWTPs), particularly
for large systems. It provides physical straining of solids as water moves through filter beds containing
sand and other media (e.g., anthracite). Downflow systems are the most common type, but continuous
backwashing upflow systems (e.g., Dynasand®) are becoming more popular. Granular filters come in a
variety of configurations based on depth, media, driving force (e.g., gravity or applied pressure through
pumping), and method of flow control.
Advanced filtration technologies have been developed to provide enhanced performance and, in
some cases, reduced footprint requirements. Some are adaptations of conventional granular media
filtration. For example, the pulsed bed filter intermittently pulses air through a shallow sand bed to
bring clean sand to the top and bury trapped solids, thereby making use of the full depth of a single filter
media. The traveling bridge filter is a shallow, low-pressure head filter with an overhead traveling bridge
assembly for collecting backwash water. Membrane filters have smaller pore sizes and can remove
solids to very low levels, but they require higher pressures to operate. Specialty media filters have been
designed specifically to remove additional phosphorus. See Chapter 6, Section 6.5 for a summary of
available effluent filtration technologies.
Several publications provide detailed design guidance for effluent filtration (e.g., WEF and ASCE
2010; WEF 2006; Tchobanoglous et al. 2003). The purpose of this chapter is to provide key design
factors that affect the performance of effluent filtration for the purposes of removing TP to low effluent
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levels. Many states also have design criteria for filtration processes; designers should carefully review
state requirements and guidelines. Because of the inherent variability in influent solids to be filtered,
designers should strongly consider pilot testing to establish the design parameters for full scale effluent
filtration (Tchobanoglous et al. 2003).
A new tool for the design of tertiary filters is computational fluid dynamic (CFD) modeling. CFD
modeling can be used to predict filter performance prior to design. It can be used to complement
laboratory or bench scale tests. Water Online (2009) reports that CFD modeling was used to predict
performance of a deep bed granular media filtration technology for an upgrade to the Blue Plains WWTP
in Washington, D.C.
11.2 Selection of Filtration Technology
Selection of a filtration process depends on many factors including effluent goals, available
space, operational considerations, reliability, and costs. Chapter 8 of this design manual provides
general guidelines for selecting candidate treatment processes for plant upgrades. Some key
considerations for selecting a specific tertiary filtration technology are highlighted below.
• Desired Effluent TP. When operated properly, most conventional granular media filters can
achieve consistent effluent levels of less than 0.1 milligrams per liter (mg/L) TP (as long as
soluble phosphorus has been removed to less than that concentration) and less than 5 mg/L
TSS1. Combination technologies, membranes, and some emerging technologies can achieve
additional TP removal. Technologies that can potentially achieve extremely low effluent TP
levels of less than 0.05 to 0.01 mg/L include the following (Bott et al. 2007; Barnard et al.
2008; EPA Region 10 2007):
- Two-stage filtration.
Enhanced clarification and tertiary filtration. Enhanced clarification may include high
rate ballasted sedimentation or sludge blanket clarification.
Membrane filtration (microfiltration or ultrafiltration).
Specialty filters (e.g., an iron oxide coated sand filter).
As noted elsewhere in this design manual, removing phosphorus through chemical/physical
processes requires a chemical addition step to convert soluble reactive phosphorus to a
particulate or colloidal form. This conversion is typically accomplished by adding a metal salt
such as alum or ferric chloride and requires proper mixing. See Chapters 3 and 9 of this
design manual for details about the theory of phosphorus removal by chemical addition and
the design of chemical feed and mixing systems.
• Mode of Operation. Operational considerations are important when evaluating filter
technologies. Some technologies, such as the continuous upflow backwashing filters and
1 Although these effluent concentrations have been observed in the field, technology performance depends on a
number of factors and low limits are may not be consistently achievable in all cases. See Chapter 6 for full
discussion of technology performance.
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the traveling bridge filter, do not need to be taken out of service for backwashing. Shallow
bed systems such as pulsed bed and traveling bridge filters require a lower flow rate and
generate less backwash water compared to conventional granular media systems. The
traveling bridge filter produces a relatively constant amount of backwash when the filter is
in operation, which may eliminate the need for storage of the backwash water before it is
introduced to the head of the treatment plant.
• Size Constraints. The surface area required for filtration depends on the hydraulic loading
rate. Cloth, disk, and membrane filters operate at higher loading rates and thus, have
smaller footprints than traditional granular media filters. Pulsed bed and traveling bridge
filters have large footprints but have relatively shallow filter beds and thus, may require less
pumping for regular operation.
• Flexibility in Operations. The total filter area should be provided in two or more units,
allowing for continued operation if one filter needs to be taken off-line for backwashing or
maintenance.
• Costs. Costs can be one of the most important factors in selecting a WWTP upgrade option.
Cost estimates for comparing filtration technologies should be life cycle costs and should
include all capital costs for construction, as well as operations and maintenance costs. For
tertiary systems, it is important to weigh the additional capital and operating costs for
advanced systems with more operator training and time required for less advanced,
manually operated and adjusted systems. Cost estimates vary widely from facility to facility;
refer to the USEPA Municipal Nutrient Removal Technologies Reference Document (2008b)
for an evaluation of cost factors that should be considered for tertiary filtration systems.
11.3 Granular Media Filters
Granular media filters remove solid material from wastewater by passing the water through a
filter bed containing media such as sand or anthracite. Suspended solids are removed primarily through
physical straining—particles larger than the pore spaces are removed mechanically, while solids smaller
than the pore spaces can be removed through chance contact. As solids accumulate within the filter
media, flow will decrease and headless will increase. Filters are periodically taken out of service and
backwashed, typically on the basis of effluent quality or measured headless, to remove the accumulated
solids from the media.
The filter medium can be a single type (typically sand), dual media (commonly sand and
anthracite), or multimedia (e.g., sand, anthracite, and garnet or ilmenite). The depth of the filter is
classified as conventional (typically 2 feet) or deep bed (typically 4 to 6 feet). The filter medium
commonly rests on a layer of gravel, which then rests on a filter underdrain system, although a direct
media retention drain can be used which does not require gravel. The deeper the bed, the longer the
filter can typically operate without backwashing, but a higher degree of pumping is required.
Downflow filters are equipped with an inlet channel to disperse the water evenly over the filter.
They have a backwash system with control valves to stop the influent flow, a pump to force water up
through the bed to expand the media, and collection troughs above the top of the media to collect the
backwash water. Clean water for backwashing is usually pumped from a filtered water clearwell or
Nutrient Control Design Manual 11-3 August 2010
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provided by gravity from an elevated tank (Tchobanoglous et al. 2003). A surface wash may be used
prior to backwashing to break up the solids that accumulate on the top of the filter.
As effluent filtration has become more common for wastewater treatment, designers and
manufacturers have modified the conventional downflow technologies in different ways to improve
performance. Traveling bridge filters and upflow continuous backwashing filters do not need to be taken
off-line for cleaning. Pulsed bed filters make efficient use of the entire depth of a single medium. As
noted in the previous section, designers may want to consider one of these modifications for improved
operations and performance depending on plant conditions and treatment goals.
The following are key design factors for removing particulate phosphorus from wastewater and
are discussed in this section:
• Influent water quality
• Media specifications
• Filter loading rate
• Headless
• Backwashing
• Flow control
Additional information and guidelines for design are provided in standard industry references
such as Water Environment Federation (WEF) and American Society of Civil Engineers (ASCE) Manual of
Practice No. 8 and Tchobanoglous et al. (2003).
11.3.1 Influent Water Quality
The most important influent characteristics for designing a tertiary filter system are (1) the TSS
concentration, (2) the particle size and distribution, and (3) the floe strength. The TSS of secondary
effluent typically ranges from 6 to 30 mg/L, with corresponding turbidity values of 3 to 15 nephelometric
turbidity units (NTU). The distribution of particles tends to be bimodal, meaning that there are two
distinct peaks. One is typically between 0.8 andl.2 micrometers (u.m), and the second is between 5 and
100 u.m (Tchobanoglous et al. 2003). There are also a few larger particles that do not readily settle in
secondary clarifiers. Pilot testing is recommended to address the inherent variability in influent water
quality.
If a chemical coagulant such as alum or ferric chloride is added directly prior to the filter for
phosphate removal, the chemical should be well mixed to ensure formation of the solid precipitate.
Organic polymers can be added to increase particle size and improve filtration performance. Jar testing
should be used to determine coagulant and/or polymer dose. Pipes or conduits ahead of filters should
be designed to minimize shearing of floe particles (Ten State Standards 2004).
11.3.2 Media Specifications
In a mono-media filter, larger media will settle to the bottom and smaller media will settle at the
top during backwashing. This settling pattern results in a large proportion of solids being removed in the
first few inches of the bed, leaving the rest of the bed unused and resulting in high headless and short
filter runs. Dual and multimedia filters are more commonly designed to utilize the entire filter bed by
Nutrient Control Design Manual 11-4 August 2010
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layering the media with larger, lighter media on the top (e.g., anthracite) and smaller, heavier media
(e.g., sand and garnet) on the bottom. During the backwash, media will stratify as they settle, retaining
the desired gradation.
Important characteristics of granular filter media are density, sphericity, size, uniformity
coefficient, and depth of media. Table 11-1 lists the most common filter media along with information
on density and sphericity. Table 11-2 provides some typical values for depth, particle size, and
uniformity coefficient for different filter types.
Table 11-1. Common Filter Media and Characteristics
Media
Sand
Anthracite
Garnet
llmenite
Density (Typical),
2.65
1.6
4.2
4.5
Sphericity
0.40-0.60
0.75-0.85
0.60-0.80
N/A
Source: Tchobanoglous et al. (2003)
Table 11-2. Filter Media Depths and Particle Sizes
Media
Depth (in)
Effective Particle Size
(mm)
Uniformity
Coefficient
Conventional Single Media (Sand or Anthracite)
Sand
Anthracite
18-30
24-36
0.015-0.03
0.03-0.08
1.2-1.6
1.3-1.8
Deep Bed Single Media (Sand or Anthracite)
Sand
Anthracite
36-72
36-84
0.08-0.12
0.08-0.16
1.2-1.6
1.3-1.8
Dual Media
Sand
Anthracite
7-15
15-36
0.015-0.03
0.03-0.08
1.2-1.6
1.3-1.6
Multimedia
Garnet
Sand
Anthracite
2-6
9-20
9-24
0.008-0.024
0.015-0.03
0.04- 0.08
1.5-1.8
1.3-1.8
1.4-1.8
Source: Tchobanoglous et al. (2003)
The depth of the filter bed will depend on the media size and the desired operation. Deeper
beds can go longer between backwashes; however, they will incur larger headless and require more
vigorous backwashing. The effective particle size is defined as the 10 percent size by mass and may be
designated as di0. The uniformity coefficient is a measure of the uniformity of the size of the media. It is
defined as the ratio of the 60th percentile diameter to the 10th percentile diameter. While perfectly
uniform media is not necessary, too large a uniformity coefficient can lead to smaller particles filling in
the pore spaces between larger particles, which can cause filter clogging to happen earlier. The
uniformity coefficient should be 1.7 or less (WEF and ASCE 1998; Ten State Standards 2004). Typically
the effective size and uniformity coefficient values are used to specify granular media.
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11.3.3 Filter Loading Rate
The filter loading rate, also called the filtration rate, is the flow per unit area applied to a filter,
typically expressed in gallons per square foot per minute (gal/ft2-min) or liters per square meter per
minute (L/m2 -min). The design rate depends primarily on the floe strength and size of filter media and
typically varies between 2 and 8 gal/ft2-min (Tchobanoglous et al. 2003).
The number and area of the filters should be such that the design filter loading rate is not
exceeded when one filter is taken off-line for backwashing. It may be desirable to have flow equalization
before the filters to avoid large peaks in loading rates.
11.3.4 Headloss
The operational headless through the filters can be especially important for retrofits where a
limited amount of pressure head to the filters is available before additional pumping is required. Two
important design parameters are clean filter headless and terminal headless (i.e., maximum headless
before the filter is backwashed).
Several equations exist for headless in a clean filter. One equation popular in the water industry
is the Kozeny equation (AWWA 1999):
h = Lku/pg * (l-s)2/s3Sv2V Eq. 11-1
Where:
h = Headloss in ft
L = Depth of the filter bed in ft
k = Kozeny constant which is approximately 5 for most filter media
u. = Viscosity of water in lb*s/ft2
p = Density of water in Ib/ft3
g = Acceleration due to gravity (32 ft/s2)
e = Porosity of the media
Sv = Specific surface, the ratio of volume to surface area for a particle
V = Filtration velocity in ft/s, which is the filter loading rate divided by the total filter area
For dual and multimedia filters, the equation is calculated for each media layer and summed to get the
total. The size fractions are determined by a sieve analysis of the media. Porosity and the specific
surface depend on the characteristics of the media. The specific surface can also be denoted as 6/cpd
where cp is the sphericity and diameter. The media supplier should be able to supply values of sphericity,
porosity, and sieve analysis results for the media to be used.
The terminal headless of a granular media filter is the clean bed head loss plus the increase in
headless from removal of particulate material from the water. The design value depends on the solids
storage capacity and solids capture efficiency of the filter and is usually based on prior experience or
pilot studies (AWWA 1999).
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11.3.5 Backwash Requirements
The purpose of the backwash is to remove trapped solids from the filter. In a conventional filter,
backwash is accomplished by closing the filter effluent valve, stopping flow into the filter, and pumping
clean water back through the media from the underdrains to the surface. The backwash velocity is
increased until the bed is fluidized and expanded. The velocity of the water and the abrasion from
particle collisions causes particles to separate from the filter media. The lighter entrapped particles are
carried upward by the fluid velocity and collected in the backwash trough. The filter media settles back
into the bed once the backwash is stopped. Backwash water is typically returned to the primary clarifiers
(WEF and ASCE 2010).
The following equation can be used to estimate the expanded bed depth (Tchobanoglous et al.
2003):
Le/L = (l-a)/(l-(v/vs)0'22) Eq. 11-2
Where:
L = Unexpanded bed depth in meters (m)
Le = Expanded bed depth in m
a = Porosity
v = Backwash velocity in m/s
vs = Particle settling velocity in m/s
The particle settling velocity, Vs, is given by:
vs = g(sgp-l)d2p/18v Eq. 11-3
Where:
g = Acceleration due to gravity, 9.8 m/s2
sgp = Specific gravity of the media
dp = Diameter of the media in m
v = Kinematic viscosity in m2/s
Lg/L is important in determining the required freeboard and the distance from the top of the
unexpanded filter bed to the bottom of the backwash trough. Tchobanoglous et al. (2003) recommends
that the backwash troughs be set 2 to 6 inches above the expanded filter bed and have a minimum
freeboard of 24 inches at the upper end of the trough.
Because of the nature of wastewater, additional abrasion may be necessary to achieve proper
filter cleaning. This is especially true with anthracite media, which (because of its nature) can adsorb
organic components such as oils and greases. Additional scour can be obtained by adding air to the
backwash and/or by a surface scour. Care should be taken to minimize loss of anthracite media during
the design of the air scour system. It is important to note that deep bed filters cannot be fully fluidized
and will require air scouring to achieve proper cleaning.
Filters can be backwashed either at set intervals or based on a given headless or effluent
turbidity. They can also be operated with a combination of the two, backwashing the filter on set time
Nutrient Control Design Manual 11-7 August 2010
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intervals with an alarm set to automatically backwash the filter if turbidity or headless levels get too
high.
One way to extend the time between backwashes is pulsed bed filters. Pulsed bed filters are
proprietary shallow bed filters. They use the entire bed of the filter to capture and store solids. Pulsed
bed filters have air trapped in the underdrain system. A pulse of air is periodically released and allowed
to flow upward through the media bed. The air disrupts the filter bed and allows the solids to penetrate
further into the bed, breaking up the mat of solids that would otherwise form on the surface of the filter
and enabling longer filter runs.
Automatic Backwash Designs
Figure 11-1 shows a schematic of an upflow continuous backwashing filter. The water is
introduced at the bottom of the media and allowed to flow upward through the media. It is collected at
the top of the filter and sent to the effluent. Media are allowed to flow downward through the filter and
are collected at the bottom. An airlift pipe at the bottom of the filter collects spent media and
transports it to the top of the filter using compressed air. As the spent media and water are transported
upward by the compressed air, captured particles are scoured from the media. The media and dirty
water are discharged into a washing chamber at the top of the filter. The spent sand is allowed to settle
and is then returned back to the top of the filter. The captured particles and water carrying them are
discharged through a reject compartment as backwash water. A common trade name for this kind of
filter is Dynasand®.
Continuous backwash filters have several advantages, including capability to handle higher TSS
concentrations without binding, low maintenance due to no moving mechanical parts, low headless, and
no equalization required for backwash water. Disadvantages are the foreign objects can plug the air lift
pipe and most units are proprietary.
Nutrient Control Design Manual 11-8 August 2010
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A
A-Feed
B- Feed assembly
C- Distribution
D-Sand bed
E- Filtrate
F- Airlift pump
G-Airlift discharge
H- Reject compartment
I-Washer section
J- Filtrate weir
K- Reject weir
L- Reject line
Figure 11-1. Example of Upflow Continuous Backflow Filter (DynaSand® Filter)
Source: Courtesy of Parkson Corporation
11.3.6 Flow Control
There are several ways to control the flow rate through a conventional downflow granular
media filter. The most common methods are constant-rate filtration with a fixed head, constant-rate
filtration with a variable head, and variable-declining-rate filtration. In constant-rate filtration with a
fixed head, the filtration rate is kept constant. Pumps, weirs, or effluent control valves are used to
maintain a constant flow and constant water level in the filters as solids accumulate and the headless
increases. In constant-rate filtration with variable head, pumps and weirs are used to keep the flow
through the filter constant, but the water level above the filters is allowed to change. The filter is
backwashed when the level reaches a pre-set level. In variable-declining-rate filtration, the flow is
allowed to decrease as the headless in the filter builds up. The filter is backwashed when the minimum
design flow is reached.
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It is also possible to have pressure filters where the filter is contained in a pressure vessel, and a
high pressure is applied to drive the liquid through the filter. These filters can achieve higher filtration
rates than conventional filters and require less space. They are generally used at smaller treatment
plants but do require more maintenance and care.
11.4 Cloth Media Filters
Cloth media filters remove solids by passing water through a cloth disk. The cloth panels are
installed vertically inside a tank with water moving horizontally through the panels. Cloth filters are
generally designed in one of two configurations, although nearly all of them are proprietary and will
involve slight differences. In one configuration, two or more cloth disks are installed vertically inside a
tank. Water to be filtered is introduced between the disks and flows outward through the cloth. The
disks slowly rotate with approximately 60 to 70 percent of the surface area submerged at any given
time. The un-submerged portion is sprayed with backwash water to remove accumulated solids. The
dislodged solids are collected in a backwash trough and recycled to the head of the plant. The backwash
can be applied continuously and does not require taking the unit off-line.
Another configuration also consists of a series of vertical disks covered with cloth submerged in
a tank. In this case, flow is from one side of the tank through the cloth and into a collection tube or
header. The disks do not rotate and are fully submerged. When the pressure drop reaches a pre-
determined level, the unit is taken off-line and backwashed. Because of the nature of the cloth material,
cleaning with a jet spray is occasionally needed. Figure 11-2 shows an example of this configuration.
Overflow
Weir
Disk
Drive
Motor
Influent
Weir
Effluent
Weir
Influent
Valve
Backwash
Waste
Solids
Valve
Backwash
Valve
Solids
Collection
Manifold
Backwash/
Solids Pump
Figure 11-2. Cutaway view of AquaDisk® cloth media filter.
Source: http://www.aqua-aerobic.com/default.asp , reprinted with Permission from Aqua-Aerobic Systems, Inc.
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Cloth filters have pore sizes on the order of 10 to 30 u.m. They can operate at a higher hydraulic
loading rate than granular media filters and require a lower backwash rate since the media do not need
to be fluidized. They have been traditionally installed in small plants (5 to 10 MGD) but have recently
been used at larger installations (USEPA 2008b).
Lin (2007) reported results from two pilot tests of AquaDisk® cloth media filters with respect to
phosphorus removal. Design flows were approximately 2 and 6 MGD, respectively, and hydraulic loading
rates to the filter ranged from 3 to 6.7 gpm/ft2. Prior to filtration, a two-stage system provided rapid mix
of ferric sulfate and polymer followed by flocculation. Pilot testing results showed that both ferric
sulfate and polymer addition were key to TP removal to concentrations of less than 0.3 mg/L
A new filtration technology, the Ultrascreen® Microfilter was recently tested at the Southern
Water Reclamation Facility in Palm Beach, FL The unit is a cloth disk filter but unique in that the disks
are mounted at less than 90 degrees, resulting in tangential filtration. Dalton et al. (2008) observed that
it can operate at higher hydraulic loading rates (up to 12 gpm/ft2) compared to traditional disk filters.
The authors did not evaluate effluent TP levels but observed very low effluent turbidity of 2 NTU.
11.5 Low-Pressure Membranes2
Membrane filtration removes particles by forcing water through a porous membrane.
Membranes can be classified by the pressures used to operate them, the pore size, or the mechanism by
which separation is achieved. Microfiltration (MF) and ultrafiltration (UF) are considered low pressure
membranes with operating pressures up to about 90 pounds per square inch (psi). They have relatively
well defined pores and generally operate based on size exclusion, with particles larger than the average
pore size being rejected and those smaller passing through the membrane. MF and UF systems used for
wastewater treatment have pore sizes that typically range from 0.04 to 0.2 microns. Nanofiltration (NF)
and reverse osmosis (RO) are high pressure membranes, operating at pressures up to 1,200 psi, and are
typically used for water reuse applications. Table 11-3 shows the characteristics of different membrane
types.
Table 11-3. Membrane Characteristics
Membrane Type
Low Pressure
High Pressure
Microfiltration (MF)
Ultrafiltration (UF)
Nanofiltration (NF)
Reverse Osmosis (RO)
Pore Size (microns)
0.1-10
0.01-0.1
0.001-0.01
< 0.001
Operating Pressure (psi)
1-30
3-80
70 - 220
800 - 1200
Source: WEF (2006)
In recent years, low-pressure membranes have been considered for tertiary filtration. Low-
pressure membranes can consistently achieve significant reductions in TP as long as it has been
Note that this section discusses membrane technologies for effluent filtration. See Chapter 6 for a discussion of
membrane bioreactors (MBR).
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converted to particulate and colloidal form. Significant phosphorus removal to less than 0.01 mg/L has
also been reported for pilot and demonstration studies (Bott et al. 2007; see Section 11.5.6 for more
information). WEF (2006) reports that based on results of a 2003 survey of full-scale systems, the largest
use of low-pressure membranes is to pre-treat prior to high-pressure membrane systems such as
reverse osmosis (RO) for the purposes of reuse.
This section highlights key design parameters for low-pressure membrane systems for the
purposes of tertiary filtration. Designers should refer to the WEF publication, "Membrane Systems for
Wastewater Treatment" (2006) for additional information.
11.5.1 Membrane Material
Membranes used in wastewater processes are generally made of cellulose or other organic
polymers. Although inorganic membrane materials exist, they are generally expensive and have not
found wide use in municipal wastewater processes. Elements to consider in selecting a membrane
material include thermal stability, chemical stability, cost, and chlorine resistance.
Thermal stability can be important in areas experiencing high temperatures. Some membrane
materials can degrade at high temperatures; for example, cellulose acetate membranes cannot be used
at temperatures higher than 86°F. Different types of membrane materials can be degraded by different
types of chemicals such as aromatic hydrocarbons, acids, and chlorine. Table 11-4 shows the advantages
and disadvantages of various membrane materials.
Nutrient Control Design Manual 11-12 August 2010
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Table 11-4. Advantages and Disadvantages of Membrane Materials
Material
Cellulose Acetate
Polyamide
Polypropylene
Polysulfone
Polytetraflouroethylene
Polyvinylidene Fluoride
Advantages
Inexpensive
Easy to fabricate
Thermally stable to 122° F
Stable for pH 3 to 11
Moderate thermal stability
Thermally stable to 167° F
pH range of 1 to 13
Good chlorine resistance
Resistant to aliphatic hydrocarbons,
fully halogenated hydrocarbons,
alcohols, and acids
Very hydrophobic
Excellent organic resistance
Excellent stability with acids, bases, and
solvents
Stable over all probable temperatures
Autoclavable
Good solvent resistance
Disadvantages
Temperature limited to below 86°F
pH limited to between 3 and 6
Poor mechanical stability
Biodegradable
Limited chlorine resistance
Chlorine sensitive
Less chemical-resistant than
polytetrafluoroethylene
Elongated pores
Chlorine sensitive
Poor resistance to aromatic
hydrocarbons, ketones, ethers, and
esters
Low pressure limits (25 - 100 psi)
Pore sizes below MF unavailable
Expensive
Less chemically resistant than
polytetrafluoroethylene
Pore sizes below UF unavailable
11.5.2 Membrane Configuration
Membranes are available in a number of configurations. Individual membranes can be
constructed in flat sheets, hollow fibers, tubes, and spiral wound configurations.
Flat sheet membranes are secured within supporting plates and placed in a housing parallel to
many other flat sheets. Water is filtered as it passes through the channels between the plates.
Hollow fiber membranes consist of small diameter hollow fiber bunched together in bundles
consisting of hundreds or even thousands of fibers. Water to be filtered is introduced into the inside of
the fiber and flows to the outside. Hollow fiber membranes are popular because the small diameter
allows for a large number of membranes to be grouped together with a much greater surface area per
volume ratio than other membrane configurations. They also require lower pressure to drive the
filtration. This configuration can lead to lower life cycle costs.
Tubular membranes are similar to hollow fiber membranes except that the size of the hollow
tube is about an order of magnitude greater. Hollow fibers are typically on the order of a few
millimeters in diameter, while tubular membranes are on the order of centimeters. The larger size of
tubular membranes makes them easier to clean and better for waters with higher solids content.
Spiral wound membranes consist of a flexible permeate spacer placed between two membrane
sheets, which are then wrapped in a spiral configuration around a perforated pipe. The membrane is
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sealed on all sides except the side exposed to the perforated pipe. Permeate is introduced along the
spacer parallel to the pipe and flows through the membranes into the perforated pipe.
Membranes can either be housed in a pressure vessel or submerged directly in an open channel
or chamber. Flat sheet membranes and hollow fiber membranes can be used in open channels or
pressure vessels. Spiral wound and tubular membranes are used in pressure vessels.
Flow through a membrane can be of two types: cross flow and direct feed flow. In cross flow
mode, water flows tangential to the membrane. Water that does not pass through the membrane is
collected at the end of the unit and recirculated with additional feed water. Direct feed flow requires all
the water to pass through the membrane on the first pass.
11.5.3 Process Considerations
Process considerations for designing membrane systems include pretreatment, backwashing,
cleaning, and integrity testing.
Solids, filamentous or fibrous material, metals, microbial growth, and some organics can cause
fouling of membranes. Fouling can lead to increased cleaning frequency and shorter membrane life
leading to lower flow rates and higher costs. Various pretreatment facilities can be installed to reduce
fouling and increase membrane life. Straining should be done to remove filamentous material prior to
the membranes. Typically strainers with an opening of 2 mm or less are used. Granular media and
cartridge filters are also used to remove larger particulates prior to the membranes. For NF and RO
systems where solids removal is more important, low-pressure membranes such as MF or UF are often
installed as a pretreatment step.
Membranes are backwashed to remove solids that have accumulated on the feed side of the
membrane. Backwashing is accomplished by taking the membrane out of service and reversing the flow.
Backwashes are more frequent with membrane applications than conventional filtration, with typical
periods between backwashes on the order of 15 minutes to an hour. Backwash is typically initiated
based on changes in transmembrane pressure (IMP). As with conventional filtration, sufficient
membrane units should be in place so that no one unit is overloaded if one membrane is off-line for
backwashing or cleaning. Backwash water is usually returned to the head of the treatment plant. WEF
(2006) reports that the total backwash volume is between 2 and 10 percent of average daily flow.
Although backwashing can be very effective in removing most contaminants, it cannot remove
all contaminants or compounds embedded in the membrane itself. For this reason, most low-pressure
membrane technologies come equipped with chemical cleaning capabilities, or clean-in-place (CIP)
systems. Procedures for cleaning may include a prolonged backwash with chemical addition or
immersion in a chemical bath. Cleaning chemicals include acids, bases, and surfactants. Designers should
consult the manufacturer to ensure use of chemicals suitable for the membrane material. CIP chemicals
will require a storage tank, mixing system, and a chemical feed pump. UF and MF membranes are
typically cleaned once every three to four months.
Integrity testing of membranes is not generally required for wastewater systems; however, it is
recommended to ensure that none of the membranes are damaged. Integrity testing can be indirect or
direct. Indirect methods measure water quality parameters such as turbidity or particle counts to
determine if the membrane is not functioning properly. They are not as sensitive as direct methods but
Nutrient Control Design Manual 11-14 August 2010
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do not require taking the membrane off-line. Direct methods generally are based on filling the permeate
side of the membrane with air. If the membrane is wet, increased pressure will be required to push
water through the membrane due to the surface tension within the pores. The pressure decay test,
diffusive air flow test, bubble point test, and sonic testing are different methods that can detect leaks in
a membrane. The frequency of testing is a balance between the inconvenience of taking a unit off-line
for testing and the need for ensuring effluent quality.
11.5.4 Pressure Drop
Pressure drop across the membrane is important for determining the required energy necessary
for the filtration process. For membranes in a cross flow configuration, the transmembrane pressure
(TMP) can be calculated using the following equation:
Ptm = (Pf+Pc)/2-PP Eq. 11-4
Where:
Ptm = Transmembrane pressure in bars
Pf = Pressure of the feed stream in bars
Pc = Pressure of the concentrate stream in bars
Pp = Pressure of the permeate stream in bars
For membranes in a direct feed configuration, the equation is:
Ptm = Pf-Pp Eq. 11-5
The transmembrane pressure in combination with the required flux across the membrane will
determine the required membrane area.
11.5.5 Flux Determination
Flux is the flow per unit area across a membrane, typically reported in Iiters/meters2-hour
(L/m2 hr) or gpd/ft2. The membrane filtration system can be operated in one of three ways: (1) the
pressure can be kept constant and the flux allowed to vary, (2) the flux can be kept constant and the
pressure allowed to vary, or (3) both pressure and flux can be allowed to vary. Nearly all municipal
membrane systems operate according to the first, on a constant flux basis. One study, however,
indicated that allowing both pressure and flux to vary may be the most effective mode of operation
(Bourgeousetal. 1999).
The flux is related to flow and area by the equation:
QP=FWA Eq. 11-6
Where:
Qp = Permeate flow rate in m3/s
Fw = Transmembrane flux rate in m/s
A = Membrane area in m2
Nutrient Control Design Manual 11-15 August 2010
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The flux can be related to the pressure drop by the equation:
FW = AP(PT-PO)
Eq. 11-7
Where AP is the membrane permeability coefficient in L/m2*h/kPa, PT is the transmembrane pressure in
kPa or psi, and P0 is the osmotic pressure of the feed solution in kPa or psi.
Kw will vary depending on the temperature, membrane characteristics, and influent water
characteristics. In general during design, an operating pressure will be chosen and the flux determined
for the desired flow rate. The area of membrane required can then be calculated. The process can be
iterated if the given area is not feasible. Typical operating pressures and fluxes can be obtained from the
manufacturer for a given membrane. Immersed membranes typically operate between 16 and 24
gpd/ft2. Pressurized inside and outside, they can operate at high flux values of between 30 and 35
gpd/ft2 (WEF 2006).
11.5.6 Performance Data
Table 11-5 summarizes results from recent pilot studies to evaluate the removal of phosphorus
using low-pressure membranes as tertiary filters.
Table 11-5. Phosphorus Removal Reported From Membrane Pilot Studies
Location
Spokane,
WA
Coeur
D'Alene, ID
Type of Study
and Test Period
Demonstration
pilot, Aug to Dec
2007
Pilot, May to Oct
2006
Membrane Tested
Submerged hollow
fiber UF with
nominal pore size
of 0.05 by Koch
Immersed hollow
fiber UF with
nominal pore size
of 0.04. ZeeWeed
500 system by
Zenon
Configuration
Two 30 m2
modules and
tank
Not reported
Operating
Flux
(gpd/ft2)
20-34
Not reported
Effluent TP
(mg/L)
0.3 to 1.7,
(Avg = 0.7)
0.024 to
0.067
Comments
Removal was
highly
dependent on
alum addition
prior to filter
Excursions
caused by
problems with
alum feed or
inadequate pH
control
Sources: Stahyamoorthy et al. (2008); Benisch et al. (2008)
11.6 Emerging Filtration Technologies for Phosphorus Removal
11.6.1 Two-Stage Filtration
Two filters in series can provide enhanced solids removal and process control. EPA Region 10
(2007) reported excellent phosphorus removal for the Breckenridge, the Snake River, and the Alexandria
WWTPs using a two-stage treatment process with chemical addition and tertiary settling prior to
filtration. A proprietary two-stage filter, the Dynasand D2 filtration system by Parkson, has been tested
at several facilities and is reported to achieve extremely low effluent TP levels when combined with
chemical addition (EPA Region 10 2007; Bott et al. 2007). Benish et al. (2008) reported operational
Nutrient Control Design Manual
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problems, however, during pilot testing that resulted in effluent TP levels averaging 0.084 mg/L
compared to the target effluent limit of 0.050 mg/L.
The Dynasand D2 filter consists of two continuously cleaning upflow granular media filters in
series (see Figure 11-3 for a typical configuration). The first is a deep bed filter containing large sand
grains. It employs a proprietary process called continuous contact filtration (CCF) where coagulation,
flocculation, and solids removal occur within the filter bed. The second filter is a polishing filter and has
smaller sand grains and a shallower bed. The water is introduced into the bottom of the sand bed
through a number of radial arms. Particulates are removed as the water flows upward through the bed.
Sand media is continuously drawn into the airlift pipe at the bottom of the filter where it is scoured and
returned to the top through a central assembly. The reject rate reported by the manufacturer is 0.5
percent of feed flow, although higher rates have been reported (EPA Region 10 2007).
Alum
Alum
Secondary 1
Dynasand
Deep Bed
Filter
Reject
Water
scour \
Water
Dynasand
Standard
Bed
F liter
Final
Effluent
\ ^filter
scour
water
Figure 11-3. Parkson Dynasand D2 advanced filter system.
Source: USEPA 2008b, Figure2-18.
11.6.2 Iron Oxide Coated Media
The Blue PRO™ process is a continuous upflow backwashing filter designed specifically to
remove phosphorus to extremely low effluent levels through adsorption of precipitated phosphorus.
The filter was originally developed at the University of Idaho and has been tested at WWTPs in Idaho
and Florida. Skid-mounted systems can treat between 5 and 100 gpm, and concrete slab units can treat
up to 1 MGD. USEPA (2008a) reports that this particular system is not suitable for plants treating more
than 10 MGD due to the sheer number of units that would be required.
The filter media (sand) is coated with a hydrous ferric oxide coating, which enhances
phosphorus removal through adsorption. A ferric salt is added prior to the filter to aid in coagulation and
to replace the ferric coating, which is abraded from the sand during operation. Water flows up through
the filter while the sand travels down. An airlift tube at the bottom of the filter carries the sand upward.
Turbulence from the compressed air dislodges accumulated iron and phosphorus along with any solids.
The iron, phosphorus, and particles are wasted, while the clean sand is deposited on the top of the bed.
The Blue PRO® filter runs continuously and does not need to be taken off-line for backwashing. It has an
Nutrient Control Design Manual
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added advantage of recycling material instead of accumulating phosphorus in sludge where it is
susceptible to secondary release during processing.
See Figure 11-4 for a typical two-stage configuration. Results from pilot studies are summarized
in Table 11-6.
Moving Bed Filter 1
Moving Bed Filter 2
Seconds
Effluent
i
Fe
iry
,p. Pre-re
3+
\ ] \
ictor 1 — ^.
\
\l
/
/ ^ D_ *_ o i
s /
\ /
\ ••
Fe*
i
/ Effluen
i ^
•
Reject
Reject
Figure 11-4. Blue-PRO® process.
Source: USEPA 2008b, Figure 2-20.
Table 11-6. Pilot Test Results for the Blue Water Blue PRO System
Location
Hayden Wastewater
Research Facility,
Hayden ID
Coeur D'Alene, ID
Eastern Florida (test
results reported by
manufacturer)
Flow
(MGD)
0.25
0.0144(10
gpm)
Duration
of Study
3 months
(Dec05-Feb
06)
2 months
(Nov- Dec
07)
Influent TP
(mg/L)
6mg/L
4 mg/L
Effluent TP
(mg/L)
Avg of 0.011
0.013-0.034
(Avg of 0.021)
All results<
0.010
Operational Factors
Fe dose = 5 mg/L
Fe dose = 38 mg/L
Optimized Fe dose was key to
performance
Sources: Newcombe et al. (2008); Benisch et al. (2007); Blue Water Technologies (2008)
11.7 References
AWWA. 1999. Water Quality & Treatment: A Handbook of Community Water Supplies, 5th Edition.
Letterman, R.D., Technical Ed. McGraw-Hill, Inc.
Barnard, J.L, H. Phillips, B. Sabherwal, C. deBarbadillo. 2008. Driving Membrane Bio-Reactors to Limit of
Technology. Presented at WEFTEC 2008.
Benisch, M., D. Clark, Neethling, J.B., H.S. Fredrickson, A. Gu. 2007. Can Tertiary Phosphorus Removal
Reliably Produce 10 ug/L?: Pilot Results from Coeur D'Alene, ID. Presented at Nutrient Removal 2007.
WEF.
Nutrient Control Design Manual
11-18
August 2010
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Blue Water Technologies. 2008. Blue Pro Pilot Project Report: Phosphorus Removal from Wastewater
Located at a Municipal Wastewater Treatment Plant in Florida. Blue Water Technologies, Inc. Hayden,
Idaho.
Borgeous, K., G. Tchobanoglous, and J. Darby. 1999. Performance Evaluation of the Koch Ultrafiltration
Membrane System for Wastewater Reclamation. Center for Environmental and Water Resources
Engineering. Report No. 99-2, Department of Civil and Environmental Engineering, University of
California Davis, CA.
Bott, C.B., S. N. Murthy, T. T. Spano, and C.W. Randall. 2007. WERF Workshop on Nutrient Removal: How
Low Can We Go and What is Stopping Us from Going Lower? Alexandria, VA: WERF.
Caliskaner, O. and G. Tchobanoglous. 2005. Modeling Depth Filtration of Activated Sludge Effluent Using
a Compressible Medium Filter. Water Environment Research. 77(7): 3080.
Caliskaner, O. and G. Tchobanoglous. 2008. Development and Solution of Depth Filtration Equation.
Presented at WEFTEC 2008.
Chalmers, R.B., M. Patel, B. Dunnivin, and D. Cutler. 2008. Orange County Water District's Groundwater
Replenishment System is Now Producing Water. Presented at WEFTEC 2008.
Dalton, D., D.E. Loy, M.G. Schill. 2008. Demonstration of "Dynamic Tangential Filtration" at Palm Beach
County to Produce High Quality Reuse Water. Presented at WEFTEC 2008.
EPA Region 10. 2007. Advanced Wastewater Treatment to Achieve Low Concentration of Phosphorus.
EPA Region 10. EPA 910-R-07-002. Available online:
http://yosemite.epa.gov/rlO/water.nsf/Water+Qualitv+Standards/A WT-
Phosphorus/$FILE/AWT+Report.pdf
Hendron, L., M. Laquidara, S. Sathyamoorthy, G. Shrope, Gale Olrich, and J. Barr. 2008. Water Reuse as
Part of the City of Spokane's Integrated Approach to Reduce Phosphorus Discharge. Presented at
WEFTEC 2008.
Lin, H., D.J. Binder, and L.W. Johnson. 2008. Effect of Particle Removal by OptiFiber® PA-13 Nylon Pile
Media on Particle Size Distribution and Correlation between Turbidity and Total Suspended Solids.
Presented at WEFTEC 2008.
Moller, G. 2006. Absolute (1000-fold) Phosphorus Removal: Performance, mechanisms and engineering
analysis of iron-based reactive filtration and coupled CEPT at the Hayden, Idaho WWTP. University of
Idaho.
Newcombe, R.L., R.A. Rule, B.K. Hart, and G. Moller. 2008a. Phosphorus Removal from Municipal
Wastewater by Hydrous Ferric Oxide Reactive Filtration and Coupled Chemically Enhanced Secondary
Treatment: Part I - Performance. Water Environment Research. 80(3):238-247.
Newcombe, R.L., D.G. Strawn, T.M. Grant, S.E. Childers, and G. Moller. 2008b. Phosphorus Removal from
Municipal Wastewater by Hydrous Ferric Oxide Reactive Filtration and Coupled Chemically Enhanced
Secondary Treatment: Part II - Mechanism. Water Environment Research. 80(3):238-247.
Nutrient Control Design Manual 11-19 August 2010
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Olivier, M. and D. Dalton. 2002. Filter Fresh: Cloth-Media Filters Improve a Florida Facility's Water
Reclamation Efforts. Water Environment & Technology. 14(11). Alexandria, VA: WEF.
Sathyamoorthy, S., M. Laquidara, N. Anand, Z. Nikolic, S. Dukes, and A. von Gottberg. 2008. Operation of
Submerged Hollow Fiber Ultrafiltration Membrane for Treatment of Secondary Effluent. Presented at
WEFTEC2008.
Scott, S.A. and E.A. Lawrence. 2007. Pilot Study Application of Tertiary Clarification and Filtration to
Meet Proposed Ultra Low Phosphorus Discharge Limits on the Spokane River. Presented at Nutrient
Removal 2007. WEF.
Tchobanoglous, G., F. L. Burton, and H.D. Stensel. 2003. Wastewater Engineering: Treatment and Reuse.
New York, NY: McGraw-Hill.
Ten States Standards. 2007. Recommended Standards for Water Works, 2007 Edition: Policies for the
Review and Approval of Plans and Specifications for Public Water Supplies. Water Supply Committee
of the Great Lakes - Upper Mississippi River Board of State and Provincial Public Health and
Environmental Managers. Available online: http://10statesstandards.com/waterstandards.html
USEPA. 2008a. Emerging Technologies for Wastewater Treatment and In-Plant Wet Weather
Management. EPA 832-R-06-006. Available online: http://www.epa.gov/OW-
OWM.html/mtb/emerging_technologies.pdf
USEPA. 2008b. Municipal Nutrient Removal Technologies Reference Document., Volume 1 -Technical
Report. Office of Wastewater Management, Municipal Support Division. EPA 832-R-08-006. Available
online: http://www.epa.gov/OWM/mtb/mnrt-volumel.pdf
Water Online. 2009. ITT Leopold Completes Tertiary Filtration Project for Wastewater Treatment Plant
in Washington, D.C. Website accessed 9 April 2009. Available online:
http://www.wateronline.com/article.mvc/ITT-Leopold-Completes-Wastewater-Proiect-ln-N-0001
WEF and ASCE. 1998. Design of Municipal Wastewater Treatment Plants - MOP 8, 4th Ed. Water
Environment Federation and American Society of Civil Engineers. Alexandria, VA: WEF.
WEF and ASCE. 2010. Design of Municipal Wastewater Treatment Plants - WEF Manual of Practice 8 and
ASCE Manuals and Reports on Engineering Practice No. 76, 5th Ed. Water Environment Federation,
Alexandria, VA, and American Society of Civil Engineers Environment & Water Resources Institute,
Reston, Va.
WEF. 2006. Membrane Systems for Wastewater Treatment. Alexandria, VA: WEFPress.
WERF. 2008. WERF Nutrient Challenge, Chapter 1: Tertiary Phosphorus Removal. Accessed 11 March
2009. Available online: http://www.werfnutrientchallenge.org/chapterl.asp?area=chl
Nutrient Control Design Manual 11-20 August 2010
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12. Operation and Optimization to Enhance Biological Nutrient Removal
Chapter 12 covers:
12.1 Introduction
12.2 Analysis of Existing Operations
12.3 Incorporating SCADA and other Instrumentation
12.4 Common Operational Changes
12.5 References
12.1 Introduction
Depending on the existing plant configuration and operating conditions, it may be possible for
plants to meet new performance goals by optimizing existing treatment processes, particularly if
biological nutrient removal is already being practiced. It is often more cost effective to optimize
operations than to add a new treatment process or to construct new reactor volumes.
Process simulation models can be extremely useful in optimizing the operation of an existing
wastewater treatment plant (WWTP). Operators can adjust parameters and predict their impact
without having to do extensive testing at the plant. Models can also help to identify bottlenecks at the
plant.
This chapter provides general guidelines for evaluating an existing plant and changing plant
operation to improve performance. While this chapter includes many examples of plant operational
improvements, each plant is unique, and alternative approaches may work better for a given situation.
Operators and designers should always apply the fundamental process information provided in Chapters
3 through 5 of this manual to address operational issues and improve performance.
12.2 Analysis of Existing Operations
Process optimization begins with an analysis of the performance of each treatment process at
the existing plant. The purpose of this analysis is to answer questions such as:
• What are the factors that are limiting nutrient removal?
• Where are the hydraulic bottlenecks?
• Do parallel treatment units achieve different performance levels for nutrient removal? If so,
why?
• What is causing variability in effluent quality with respect to nutrients?
• What are the causes of intermittent increases in nutrient concentrations in the plant
effluent?
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This section describes data analysis and modeling techniques that can be used to answer these
questions and ultimately improve performance of the existing WWTP.
12.2.1 Data Analysis
The following 4-step procedure is one possible approach to collecting and analyzing plant data
for the purposes of process optimization. Alternative approaches can be found in Chapter 15 of
Tchobanoglous et al. (2003) and Chapter 12 of WEF and ASCE (2006). Note that all data analyses should
be done in close coordination with plant operations staff. Plant personnel may have unique insight into
the influence of operational changes on measured parameters, and vice versa, that could be extremely
valuable for process optimization.
Note that all influent, primary effluent, secondary effluent, and plant effluent samples should be
24-hour composite samples. Grab samples will not usually provide the necessary accuracy, and results
can be very misleading.
Step 1: Compile Existing Data
Obtain as much data as possible for the influent, primary effluent, secondary effluent, and plant
effluent. See Table 12-1 for a list of recommended parameters for data analysis.
Step 2: Collect Additional Data
If the data are not sufficient to evaluate plant operations or if designers suspect that some data
may not be accurate, consider collecting and analyzing additional samples to characterize plant
performance. At least some form of special sampling is usually needed for process optimization. A
successful sampling campaign requires careful planning and execution using proper techniques for
sample collection, handling, and laboratory analysis. See Chapter 7, Section 7.4, for recommendations
regarding measurement campaigns. Below are additional considerations for sampling for the purposes
of process optimization.
• If operators suspect inconsistent performance among parallel treatment trains, each should
be sampled separately. Alternatively, intense sampling can be done for many locations
along one treatment train. Include flow measurements for each of the parallel trains.
Consider performing a flow balance for the entire treatment system as a check (include
comparison of influent and effluent flows).
• Consider interval sampling, particularly if plant personnel suspect that short-term variations
in influent quality and/or intermittent discharge of recycle streams have a significant impact
on the nutrient loading to the plant. Samplers can be programmed to segregate samples for
specific time periods (WEF and ASCE 2006). A duration of 2 to 3 days is typically sufficient
for interval sampling.
• Consider installing online monitoring equipment. If they are integrated into the process
control system, online monitors can provide the added benefit of enabling real-time
operational changes. See Section 12.3 for additional discussion of automated process
control.
Nutrient Control Design Manual 12-2 August 2010
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Table 12-1. Recommended Parameters for Data Evaluation
Parameter1
Flow
Total BOD/COD
Soluble1
BOD/COD
TSS, VSS
DO
PH
Alkalinity
Temperature
NH3-NorTKN
Nitrate
Total Phosphorus
Phosphate
ORP
VFA or rbCOD
MLSS&MLVSS
test and WAS test
Location(s)
Influent, effluent, flow splits,
recycles
Influent, primary effluent,
effluent, anaerobic & anoxic zone
effluents
Primary effluent, secondary
effluent, final effluent
Aerobic, anaerobic, and anoxic
zones (multiple locations
recommended)
Influent, mixed liquors, effluent
Influent, primary effluent, mixed
liquor supernatants, effluent
Influent, mixed liquors of
reactors, effluent
Influent, primary effluent, reactor
mixed liquors, secondary
effluent, effluent
Influent, reactor mixed liquors,
secondary effluent, effluent
Influent, primary & secondary
WAS, plant effluent
Reactor mixed liquors, primary &
secondary effluents, effluent
Anaerobic & anoxic reactor mixed
liquors
Influent, primary effluent
MLSS & MLVSS: well-mixed
location in aeration basin
WAS: well-mixed and
representative sample from the
WAS pipe (may need composite
sample)
Rationale
Essential for developing mass balances, which are essential for a
complete understanding of the treatment system.
Can be used to evaluate substrate availability for biological
processes. High effluent BOD could indicate activated sludge
performance problem.
Important if phosphorus is removed chemically. Used to calculate ISS,
determine clarification efficiency, and determine an accurate solids
residence time (SRT) for the bacteria.
Minimum DO of 2.0 is usually needed to minimize oxygen limitation
of nitrification rates, which is important for low SRT/HRT systems.
DO should not be present in anoxic or anaerobic zones.
Should be above 6.5 and below 9.0 for biological nitrogen removal.
Low pH or wide swings in pH could mean significant industrial
component. Could affect BPR and nitrification.
If effluent is below 50, there is probable nitrification inhibition, and
process is susceptible to large pH drops as a result of nitrification or
chemical addition for phosphorus removal.
Low temperatures can significantly reduce nitrification rate. For the
typical range between 10 and 25 °C, the rate will drop by half for
every 8 to 10 °C reduction in mixed liquor temperature. Reactor
temperatures are likely to be significantly different from influent
temperature because of aeration.
Can be used to evaluate load to and performance of biological
nitrification kinetics.
A check on nitrification, and can be used with TKN to calculate
denitrification.
Used to calculate phosphorus removal efficiency by treatment
processes.
Used to determine release and uptake in reactors, release in
secondary clarifier, and phosphorus removal efficiency.
Measures the balance between oxidized and reduced compounds
present in solution. Will detect presence of significant
concentrations of oxidized compounds. Can be used for automatic
detection of excess electron acceptors (DO, nitrate, and nitrite) in
reactors.
Can be used to evaluate substrate availability for enhanced biological
phosphorus removal.
MLSS and WAS tests provide suspended solids concentrations and
can be used to determine percent phosphorus in sludge. This
information, in conjunction with aeration basin volume and WAS
flow, can be used to calculate SRT. Maintaining SRT is critical for
nitrification and, sometimes, for enhanced biological phosphorus
removal.
' BOD = biochemical oxygen demand
(5-day unless otherwise noted)
COD = chemical oxygen demand
TSS = total suspended solids
DO = dissolved oxygen
TKN = total Kjeldahl Nitrogen
VFA = volatile fatty acids
MLSS = mixed liquor suspended solids
MLVSS = mixed liquor volatile suspended solids: Inorganic suspended solids (ISS) = MLSS - MLVSS
Source: WEF and ASCE (2006)
Total BOD/COD = unfiltered BOD/COD
Soluble BOD/COD = BOD/COD of filtrate from 0.45 [im pore size filter
VSS = volatile suspended solids
ORP = oxidation reduction potential
NH3-N = ammonia Nitrogen
rbCOD = readily biodegradable COD
WAS = waste activated sludge
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Step 3: Review and Summarize Data
Review data for anomalies and possible sampling errors. This step is also called data cleaning or
data verification. See Section 7.4.2 of this design manual for recommended data checks. Mass balance
diagrams are particularly helpful, and should be developed for all loading information and for inert
fractions (ISS) through the plant. Additional samples may need to be taken as a result of these steps.
The evaluation of plant performance is facilitated by presenting data in carefully annotated
tables, graphs, and figures. Below are possible approaches to organizing plant data for analysis:
• Summarize key parameters in tabular form.
• Consider annotating a process flow diagram with flow weighted averaged data (Frank et al.
2008).
• Plot sampling results for key parameters over time to assess temporal trends and at
different locations to assess spatial trends.
• Consider using histograms to depict the proportion of measured values that fall in distinct
categories.
• Consider plotting multiple parameters measured at a specific location together on the same
plot to evaluate possible relationships.
• 3-dimensional graphs with measured data (e.g., ammonia), sampling location, and time are
an informative way to display both spatial and temporal trends. See Figure 12-1 for an
example using special sampling data from the Washington Suburban Sanitary Commission
(Frank et al. 2008).
Nutrient Control Design Manual 12-4 August 2010
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Sampling Location
(a)
(b)
Figure 12-1. Spatial and temporal profiles of ammonia.
Source: Frank et al. (2008), Figure 6. Reprinted with permission from Proceedings of WEFTEC®.2008, the 81st Annual Water
Environment Federation Technical Exhibition and Conference, Chicago, Illinois, October 18-22. Copyright ©2008 Water
Environment Federation, Alexandra, VA., www.wef.org.
Step 4: Evaluate Relationships between Key Parameters
Statistical analyses can provide very useful information. Consider linear regression analysis to
evaluate the relationship between water quality and operational parameters.
12.2.2 Use of Process Simulation Models
Process simulation models, particularly if they have been calibrated to dynamic operating
conditions, can be extremely useful tools for process optimization. Designers can use models to
simulate many different operating scenarios and analyze results to determine if nutrient removal
efficiencies can be improved. For example, designers can simulate changes in aeration rates to gauge
the effects on nitrification, or simulate increased wastage rate from the secondary clarifier to determine
the effect on SRT. Dynamic modeling can also be used to check the plant's ability to treat peak loads.
For example, dynamic modeling of the Beenyup plant in Western Australia showed that the nitrification
process was limited by the capacity of the diffusers to meet the peak air demand (Third et al. 2006). In
addition to dynamic simulations, steady state modeling can be useful in determining major process
bottlenecks by comparing simulated operating conditions with original process design capacities.
Steady state modeling can also be used to compare performance under different operating scenarios.
The usefulness of simulation models for process optimization relies heavily on proper model
development and calibration. See Chapter 10 of this design manual for guidance on model development
and calibration.
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12.3 Incorporating SCADA and other Instrumentation
Compared to grab sampling, online monitoring can give real-time information on processes
occurring within the treatment plant and is not susceptible to improper sampling, holding times, or
inconsistencies in laboratory analysis. Chapter 13 describes the types of instruments and control
systems for nutrient removal.
On the most basic level, online sensors can provide information to operators and assist them in
making operational decisions. More advanced online monitoring applications feed the measured data
into a controller. The controller can be programmed to automatically adjust a control parameter (e.g.,
diffuser setting) depending on the measured value. This set up is referred to as a "control loop" for
automated process control.
Historically, the lack of reliable sensors has been the biggest barrier to implementing automated
process control. Recent improvements in basic and advanced online sensors have made them more
reliable and more commonly used (Jeppsson et al. 2002). Automated control has many advantages
including reducing staff workload, allowing for more precise control of process parameters, and
providing automatic data logging and archiving (Tchobanoglous et al. 2003).
In the past, proprietary distributed control systems were recommended for WWTPs (WEF and
ASCE 2006). More and more facilities, however, are implementing comprehensive supervisory control
and data acquisition (SCADA) systems. SCADA systems encompass process control; data acquisition;
supervisory control; distributed alarming; historical collection, display, and analysis; information
systems; and security. See Chapter 13, Section 13.8 for additional discussion of SCADA functionality.
12.4 Common Operational Changes
12.4.1 Adjust SRT
Nitrification is highly dependent on sufficient SRT in the aerobic basin. Denitrification can be
limited by the anoxic SRT. Biological phosphorus removal can also wash out at low temperature-low SRT
combinations. Higher SRTs are needed for colder temperatures and vice versa.
Capital improvements to increase SRT include constructing additional basin volume; installing a
fixed film media in the basin, such as integrated-fixed film media (IFAS); or using membrane bioreactors
(MBRR). Operational improvements can sometimes be implemented at a much lower cost. Operating
changes that can be used to increase SRT include:
• Increase the MLSS/MLVSS concentration in the aeration basin. However, this is limited by
the solids separation process.
• Incorporate swing zones to utilize the maximum available aeration volume or the maximum
anoxic volume during peak loadings. This increases the aerobic or anoxic SRT.
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12.4.2 Adjust Aeration Rates
Tapered aeration with lower DO levels at the end of the aeration basin can promote
simultaneous nitrification-denitrification (SNdN). SNdN can also be increased by cycling the air on and
off. It can be optimized by automating the DO level for the former case, or the on-off cycling times for
the latter case, via automatic monitoring of ammonium and nitrate using probes and feedback to a
control system. Tapered aeration also optimizes the utilization of DO and the need to transfer DO,
resulting in the reduction of energy requirements. If anoxic reactions are increased, the amount of WAS
produced also is reduced.
12.4.3 Add Baffles to Create High Food to Microorganism (F/M) Conditions
The performance of existing continuous flow, activated sludge BNR systems can generally be
improved by using plug flow rather than complete-mix configurations, and the plug flow units can be
improved by using baffles to divide the anaerobic, anoxic, and aerobic zones into two or more sections
each. The use of multiple sections (2 or 3) in anaerobic zones is particularly beneficial because it
increases the substrate concentrations surrounding the activated sludge microorganisms in the first
section. This increases the rates of reaction within the section, which reduces the total volume needed
for the zone. Additionally, the first section protects the remaining section(s) from influent nitrates and
DO, which helps ensure the development of anaerobic conditions within the zone and provides a safety
factor against nitrate and DO excursions. Similar statements can be made regarding the anoxic zone
where the purpose of baffling is to provide protection from the recirculation of unusually high DO
concentrations, thus ensuring that anoxic conditions are obtained within the zone in addition to
obtaining higher rates of reaction in the first section.
There is less justification for the use of baffles in the aerobic zone because (1) the phosphorus
removal rates of reaction are fast relative to the nitrification rates, (2) high DO concentrations increase
both reactions, and (3) there are no other electron acceptors that interfere with the desired reactions.
Note that soluble substrate is usually near effluent concentrations when the flow enters the aerobic
zone if a full BNR system (i.e. with anaerobic, anoxic, and aerobic zones) is used. Baffling for a small
aerobic effluent section could be used to reduce the DO concentrations recycled to the anoxic zone and
entering the secondary clarifiers, the latter of which would reduce the DO recycled to the anaerobic
zone in the return activated sludge (RAS). However, adequate reduction of DO in the aeration basin
effluent can generally be accomplished by simply reducing the amount of air delivered to the last 5
percent of the aeration basin.
Baffles should always be designed to prevent backmixing. Because the water surface elevation
in aerated zones are higher than those in non-aerated zones, backmixing of DO and nitrates can cause
floating scum. Underflow baffles with tops that extend above the water surfaces are excellent
protection against backmixing. A minimum forward flow of 1 foot per second is recommended through
the underflow openings to prevent backmixing during low flow periods.
12.4.4 Change Aeration Settings in Plug Flow Basins
There are many instances where enterprising operators at conventional plants with plug flow
aeration basins have achieved 80 percent or more phosphorus removal with less than 1 mg/L total
Nutrient Control Design Manual 12-7 August 2010
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phosphorus (TP) in the plant effluent by turning off air or mixers. Switching off aeration at the inlet side
of the basin system will create an initial anaerobic zone. As long as there is enough readily
biodegradable chemical oxygen demand (rbCOD) in the plant influent, Phosphate Accumulating
Organisms (PAOs) will grow and remove phosphorus. It is important to ensure that the processes still
have sufficient SRT, particularly for nitrification in the aerated portions of the basin.
The Piney Water, CO, plant is a 5-stage Bardenpho plant with no primary sedimentation and few
VFAs in the influent, which resulted in little phosphorus removal. By switching off a mixer in one of the
anaerobic zones, sludge settled to the bottom and fermented, which supplied the VFAs for reducing the
orthophosphate to less than 0.2 mg/L. A similar operation at the Henderson, NV, plant in a
Johannesburg Process (JHB ) had the same effect.
Some plug flow aeration plants succeeded in reducing phosphorus to below 1 mg/L by turning
off aeration at the feed end of the plant, such as the Blue Lakes and Seneca plants operated by the
Metropolitan Council Environmental Service in Minnesota and the St. Cloud, MN, plant.
12.4.5 Minimize Impact of Recycle Streams
Discharge of supernatant from sludge handling operations can have a significant impact on BNR
processes, particularly for medium and small WWTPs that discharge recycled flows intermittently. For
example, a WWTP treating approximately 4 MGD correlated increased effluent TP concentrations with
discharge from sludge handling facilities. Plant operations staff tracked the problem to phosphorus
release in sludge holding tanks. Improved aeration within the holding tanks prevented phosphorus
release, reduced intermittent phosphorus loading, and subsequently allowed the plant to achieve more
consistent effluent TP levels (Smith et al. 2008).
A second potential problem is increased ammonia load in return streams from sludge
dewatering operations. In some plants, dewatering is operated during the day and supernatant from
digested sludge is returned to the head of the plant in the afternoon. This coincides with a diurnal
ammonia peak in residential systems and can overwhelm nitrification processes. WEF and ASCE (2006)
report that a small tank can significantly reduce variation of influent ammonia load by storing recycle
flow from the dewatering facility. The tank can be filled during peak hours and emptied during off-peak
hours using a simple timer.
12.4.6 Reconfigure Flow through Existing Units
Some plants may find opportunities to reconfigure flow through existing basins to enhance BNR.
For example, plant operators in the City of Xenia, OH, reconfigured their existing BNR system to include
a swing zone that could be adjusted to respond to short-term variations in hydraulic and organic loading
conditions (Smith et al. 2008). This modification allowed the plant to achieve enhanced biological
phosphorus removal to effluent TP levels of 0.5 mg/L (annual average).
The West Hickman Creek Treatment Plant in Lexington, KY, reported successful conversion of
existing primary clarifiers to anaerobic zones for biological phosphorus removal (Water and Wastewater
Asia 2006). To convert the primary clarifiers, the plant operations staff installed fine screens at the
influent and converted the anaerobic digesters to aerobic sludge holding tanks. They used respirometry
techniques to assess the availability of VFAs, evaluated the hydraulics of the existing primary clarifiers,
and determined that it was sufficient for biological phosphorus removal. The plant experienced
Nutrient Control Design Manual 12-8 August 2010
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additional benefits of improved settling and clarification and reduced use of coagulant dose for chemical
polishing.
12.4.7 Increase VFAs for Biological Phosphorus Removal
The amount of readily biodegradable organic material determines the amount of biological
phosphorus removal that can be achieved at a given plant. As stated in Chapter 5, the minimum
rbCOD/TP ratio for complete biological phosphorus removal is 10 to 16 (Barnard 2006). If this minimum
is not being met, consider the following approaches:
• Check odor control practices. Odor control in the collection system can reduce anaerobic
conditions, which can reduce the formation of VFAs in the system. Consider odor control
practices that do not adversely affect VFA production such as iron addition and pH control
between 9 and 10. See Chapter 5, Section 5.3.1 for more information.
• Consider fermenting the primary sludge by increasing the depth of the sludge blanket in the
existing basin.
Although not a strictly operational improvement, addition of fermenters has been very
successful for providing substrate for biological phosphorus removal, particularly in cold climates. For
example, a complete mix tank can be installed ahead of the primary clarifier for fermentation of the
VFAs. See Section 5.3.2 of this manual for a discussion of the principles of primary or RAS sludge
fermentation. See Chapter 10 for additional design guidance.
12.5 References
Barnard, J.L 2006. Biological Nutrient Removal: Where We Have Been, Where We are Going? In
WEFTEC 2006.
Frank, K., G. Davies, N. Shirodkar, and A. Sauvageau. 2008. Full-Scale Performance Testing and
Optimization of an Old-School Process with Limit of Technology Nutrient Removal Capabilities. In
WEFTEC 2008.
Jeppsson, U., J. Alex, M.N. Pons, H. Spanjers, and P.A. Vanrolleghem. 2002. Status and Future Trends of
ICA in Wastewater Treatment-A European Perspective. Water Science Technology. 45(4-5): 485-
494.
Lopez, C, R. Wachter, C. Meyer, B. Narayanan, and C. Baer. 2008. Optimization of Biological Nutrient
Removal Systems in Response to Increased Loadings. In WEFTEC 2008.
Quast, D., P. LaMontagne, and R. Knuteson. 2008. Putting the Operator in Charge; Improving Human
Machine Interface Leads to Significant Centrifuge Operation Savings. In WEFTEC 2008.
Smith, R.C., J. Tincu, and S. Jeyanayagam. 2008. Maximizing EBPR: Going from a 'Black Box' to a 'Glass
Box'Approach. In WEFTEC2008.
Nutrient Control Design Manual 12-9 August 2010
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Swift, J., D. Bradshaw, L Pierce, W. Robinson, P. Williamson, P. Jue, J. Damitio, D. Welce, and J.
Dufresne. 2008. Full-Scale Evaluation of On-Line Ultraviolet Nitrate and Nitrite Monitoring Systems.
In WEFTEC2008.
Tchobanoglous, G., F. L. Burton, and H.D. Stensel. 2003. Wastewater Engineering: Treatment and
Reuse. New York, NY: McGraw-Hill.
Third, K.A., A.R. Shaw, and L. Ng. 2006. Application of the Good Modelling Practice Unified Protocol to a
Plant Wide Process Model for Beenyup WWTP Design Upgrade. London, UK: IWA Publishing.
Wahlberg, E.J., B. Browne, N. Fulcher, Biju George, D. Linn, L. Scanlan, D. Siler, W. Anderson, M. Daniels,
St. Rogowski, and S. Walker. 2008. Process Optimization Saves Money and Unlocks Capacity. In
WEFTEC2008.
Water & Wastewater Asia. 2006. Biological Nutrient Removal: Online Sludge Level Monitoring Supports
Efficient Nutrient Removal. Water & Wastewater Asia. January/February 2006. Available online:
http://www.waterwastewaterasia.com/WWA archive/JanFeb06/hach.pdf
WEF and ASCE. 2006. Biological Nutrient Removal (BNR) Operation in Wastewater Treatment Plants -
MOP 29. Water Environment Federation and the American Society of Civil Engineers. Alexandria, VA:
WEFPress.
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13. Instrumentation and Controls
Chapter 13 covers:
13.1 Introduction
13.2 Factors in Selecting Instrumentation
13.3 Basic Online Instrumentation
13.4 Online Instrumentation for Nutrient Control
13.5 Types of Control
13.6 Control Equipment—SCADA
13.7 References
13.1 Introduction
Nutrient removal at wastewater treatment plants (WWTPs) requires knowledge and careful
control of many key variables. Variables such as dissolved oxygen (DO) concentration, nutrient
concentrations, solids content, and flow rates are important to monitor throughout the process.
Control of variables such as DO, solids residence time (SRT), hydraulic retention time (HRT), and return
flows can be key to achieving low nutrient concentrations in the effluent. Using instrumentation to
monitor these and other variables and automatically control certain operating parameters can lead to
process optimization and more stable performance. Use of advanced instrumentation and automatic
control is strongly recommended for plants targeting very low effluent total nitrogen (TN) levels and/or
total phosphorus (TP) levels.
This chapter will begin with a discussion of factors to consider when selecting instrumentation
and control strategies. It then provides a brief description of the types of on-line instrumentation
available (basic and instrumentation developed specifically for controlling nutrient removal processes).
Following is a general discussion of the various types of logic used in control processes and the types of
equipment used.
The following technical references provide additional information on instrumentation for
nutrient removal processes:
• Chapter 7, Process Instrumentation in the Water Environment Federation (WEF) Manual of
Practice (MOP) 11, Operation of Municipal Wastewater Treatment Plants, 6th Edition (WEF
2007).
• Chapter 13, Instrumentation and Automated Process Control in the WEF MOP 29, Biological
Nutrient Removal (BNR) Operation in Wastewater Treatment Plants (WEF and ASCE 2006).
• Sensing and Control Systems: A Review of Municipal and Industrial Experiences. (Hilletal.
2002).
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• Design of Municipal Wastewater Treatment Plants - WEF MOP 8 and ASCE Manuals and Reports
on Engineering Practice No. 76, 5th Ed. (WEF and ASCE 2010)
• The Instrumentation Test Association (ITA) offering test reports, studies, and workshop
proceedings regarding online instrumentation and automation for water and wastewater
treatment applications. Information is available online at http://www.instrument.org.
• The International Water Association (IWA) Task Group on Benchmarking of Control Strategies
for WWTPs. Information available online at http://www.benchmarkwwtp.org/
It is important to note that technology often changes faster than books and large studies. Designers
should check with professional associations such as their local WEF chapter, other utilities, and
manufacturers for the latest information. Onsite testing of various brands of the different types of
instruments under consideration can provide excellent data for analysis of the specific wastewater
based on the most current instrument models.
13.2 Factors in Selecting Instrumentation
Several factors should be considered when selecting instrumentation for nutrient removal
processes.
• Online analyzers compared to offline sampling. Grab samples may be adequate for
parameters that change slowly and don't require tight control. However, much can be gained in
process control by online equipment that can provide operators with real-time measurements,
particularly for plants trying to achieve very low TN and/or TP effluent limits.
• Accuracy and reliability. Accuracy represents how close a measured value is to the true value,
whereas reliability (or repeatability) is the measure of agreement among a number of
measurements with the same input values and operating conditions. For the purposes of
control, reliability and reproducibility are more important than accuracy. WEF and ASCE (2006)
note that it is better to use a reliable, low-maintenance meter than a very accurate meter that
requires significant maintenance. The exception is pH—accurate pH readings within +/- 0-1 unit
are desirable.
• Operations and maintenance requirements. A key factor in selecting instrumentation is the
amount of operator attention required to calibrate, clean, and maintain the equipment. If staff
is not available to perform these duties, more durable and maintenance free instruments should
be chosen. Automated cleaning and calibration are available for some instruments. The
additional capital costs required for these features are often small compared to the reduced
operator requirements and risks of failure.
• Sample collection and pretreatment. Proper design and maintenance of the sample pump,
piping, and pretreatment (e.g., as required for analysis of nutrients) are key to obtaining reliable
readings. WEF and ASCE (2006) note that these components often cost more than the
instruments themselves and recommend that designers check the manufacturers' plant
installation list to learn more about successful sample collection and pretreatment systems.
Nutrient Control Design Manual 13-2 August 2010
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Note that some manufacturers avoid sample pumps and piping by offering probes which can be
installed directly in the water.
It is important to use instruments specifically designed for wastewater—in the past, attempts were
made to adapt equipment from the drinking water industry to wastewater. These attempts were largely
unsuccessful due to the high solids and microbiological populations in wastewater.
13.3 Basic Online Instrumentation
Online instrumentation has been used for many years in wastewater applications. Many plants
have online monitors for DO and total suspended solids (TSS) for basic process control. Some biological
nutrient removal (BNR) plants also use online pH and oxidation reduction potential (ORP) meters. For
the most part, the available instruments can take real-time measurements and provide high reliability.
Table 13-1 summarizes the instruments used for basic measurements and lists advantages and
disadvantages of each type.
Table 13-1. Summary of Basic On-Line Instrumentation
Analyte
Flow
TSS
Sludge blanket
monitor
DO
PH
ORP
Type of
Measurement
Mechanical
Pressure Drop
Magnetic
Reflective Sonic
Parshall Flume
Light scattering
(back scattered)
Light Adsorption
Ultrasonic
Microwave
Ultrasonic
TSS or Turbidity
Membrane
electrode
Galvanic electrode
Optical probe
Electrode
Electrode
Advantages
Accurate
Low cost
No moving parts, no wear
No pressure drop, low maintenance,
low cost
Simple, wide flow range
Better sensitivity, wider measuring
range
Less sensitive, smaller range,
inaccurate at low ranges
Insensitive to color
Insensitive to interference
Low maintenance
See TSS
Low cost
Durable, reliable
Durable, low maintenance, reliable
Indicates true oxidizing environment
(anaerobic, anoxic, or aerobic)
Disadvantages
Wear down
Highly dependent on installation,
pressure drop
High cost, inaccurate at low flow
Limited size of conduit, can't use
aggregate lined pipe, inaccurate at low
flow
Pressure drop, requires cleaning, slow
response
Needs effective cleaning system
Able to handle fouling better without
cleaning system
Fouling, background reading required
High cost, only works for high TSS
High maintenance
Interference from hydrogen sulfide,
needs frequent calibration
Higher initial cost
Fouling
Indirect measurement
13.3.1 Flow
Flow rate is an extremely important parameter for control at wastewater plants. Various types
of flow meters are available. Mechanical flow meters depend on an impeller turning as the fluid passes
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through the meter or by moving known volumes of fluid through a series of gears. Other flow meters
measure the pressure drop as the fluid passes through a flow constriction. Magnetic meters measure
the voltage generated by a magnetic field across the flow. Reflective sonic and transmissive sonic
meters can also be used. Measuring the level in a Parshall flume is a method used for open channels.
Mechanical meters can be very accurate but need to be replaced more frequently due to the
abrasive nature of wastewater. Magnetic meters will last longer as they have no moving parts and are
unaffected by the nature of the wastewater. They are expensive, however, compared to mechanical
meters and can perform poorly at low flow.
Differential pressure meters have a range of accuracy depending on how they were installed.
Improper installation can result in very inaccurate readings. They also have a limited range, with the
maximum and minimum flows needing to be within a factor of four of each other. Differential pressure
meters also cause a head loss in the flow.
Reflective sonic instruments installed in open channels have good accuracy, low maintenance,
and do not cause a pressure drop. They are limited in the size of the conduit. Reflective sonic
instruments installed outside a closed pipe are less accurate than those installed in the pipe but still do
not cause pressure drop and are a low cost and low maintenance option. They can also have problems
measuring flow in clean fluids.
Many types of flumes are available for measuring flow in open channels; however Parshall
flumes are by far the most common (WEF and ASCE 2009). Parshall flumes are very easy to use and
come in prefabricated designs; they do, however, create head loss and have a slower response time
compared to other types of flow measurement devices.
13.3.2 Total Suspended Solids (TSS)
TSS meters are based on optical scattering of light or on the difference in the velocity of
ultrasonic waves. Microwave instruments can be used for very high TSS concentrations (> 10,000 mg/L).
Light is emitted from a light source and then measured by a receptor at a given point. The two principle
types use either forward scattered light or backscattered light. Some sensors will try to improve
reliability by adding multiple beams and sensors. These multichannel instruments may also help provide
color independent analysis, as it can be more difficult to get a reliable reading in a very concentrated
dark sludge.1 Ultrasonic meters require the measurement of a background velocity in a microbial-free
sample.
TSS meters can become fouled in wastewater. Automatic cleaning systems are important for
maintaining sensor reliability. Types of cleaning systems include water or air purging, wipers, or
ultrasonic systems. Although some sensors may be able to maintain reliable functionality with only
occasional manual cleaning, it is a good idea to have automatic systems installed to ensure regular and
effective cleaning.
1 Note, however, that most light scatter and light absorption methods utilize near-infrared light sources at 860
nanometers, which removes color issues since the light source is not a white light source.
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13.3.3 Sludge Blanket Depth
Sludge blanket depth is an important parameter for optimizing settling within clarifiers.
Controlling the sludge blanket depth can be used to prevent excessive denitrification in the blanket and
the secondary release of phosphorus.
New sludge blanket monitors use sonic techniques to determine the position of the sludge
blanket. These instruments work by sending an ultrasonic pulse from the instrument down to the
sludge blanket and analyzing the time for the pulse to travel back to the instrument. Units typically are
equipped with a cleaning mechanism to prevent buildup. Other units detect the height of the sludge
blanket by multiple depth readings to determine the point of largest change in TSS or turbidity.
13.3.4 Dissolved Oxygen (DO)
DO is a key control parameter in achieving nutrient removal. It can be a good indicator of
biological activity, especially nitrification. Having poor control of DO can lead to inadequate nitrification
and phosphorus removal, disinfection problems, and high energy costs. It is estimated that tight control
of DO can save a wastewater plant between 10 and 30 percent of operating costs (WEF and ASCE 2006).
DO can be measured by membrane electrodes, galvanic electrodes, and optical DO probes.
Membrane electrodes allow diffusion of oxygen across a membrane and measure the oxygen gas at an
electrode in the inner chamber. Galvanic electrodes such as the proprietary Zullig technology, use a
galvanic current to measure the oxygen. Optical methods measure changes in light emitted by a
luminescent chemical and relate the rate of change to the DO in solution.
Membrane electrodes are reliable and easy to use but must be calibrated frequently (typically
once per week or per month depending on the manufacturer). They can also experience interference
from chlorine, iron, and hydrogen sulfide. Galvanic electrodes are more expensive but tend to be more
durable and reliable than membrane electrodes. Optical DO probes, the latest technology in DO
measurement, do not consume oxygen or electrolytes and are very reliable.
13.3.5 pH
pH is an important variable in any biological treatment process. Online monitoring can help
operators respond to changes in pH due to wastewater composition or chemical additions. It is also an
important parameter for nitrification. The nitrification process produces acid and, if there is insufficient
alkalinity present in the wastewater, will decrease the pH to a point that will inhibit the organisms
responsible for nitrification. Online monitoring of pH can prompt the operator to add alkalinity when
necessary. Also, the rate of nitrification varies with the pH, and maintaining an optimal pH (7-8)
sometimes can be used to maintain near complete nitrification at low temperatures.
All pH meters are based on potential measurement using an electrode. The meter consists of
two primary parts, a measurement electrode and a reference electrode. They can be configured as an
electrode pair or a combination electrode in one single glass body assembly, with the latter being more
common (WEF and ASCE 2006). Although interference from other ions in solution is not likely, sodium
ions can have an affect (WEF 2007). Like most electrodes, pH meters can foul and will need to be
calibrated and replaced more frequently than in water applications.
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13.3.6 Oxidation-Reduction Potential (ORP)
ORP probes are electrodes that measure the oxidizing and reducing activity in a solution, i.e.
which dominates and by how much. They can be used to detect changes in the oxidation state of the
process water and to determine if aerobic, anaerobic, and anoxic zones are being correctly maintained.
The readout of the sensor is in millivolts (mV), with positive readings indicating an oxidative state and
negative readings indicating a reductive state (WEF and ASCE 2006) when the standard hydrogen
electrode (SHE) is used as the reference. However, the standard calomel electrode is commonly used as
the reference electrode for field measurements and it has a +241 mV reading when the SHE is zero. The
calomel electrode is more robust than the SHE, and gives relatively stable readings. Operators need to
ensure that the adjustment is made to the SHE when the calomel electrode is used.
The ORP of influent wastewater typically ranges from -50 mV to -400 mV (WEF 2007) because it
is typically reduced when it reaches the wastewater treatment plant. Because ORP is a non-specific
measurement, absolute values based on industry standards are not typically used for control. Instead,
operators can evaluate ORP for sudden changes or gradual changes over time compared to other
operating parameters.
As previously noted, ORP meters are similar to pH meters in that they consist of a measuring
electrode and a reference electrode. The reference electrode is typically the same calomel electrode
used in pH instruments (WEF and ASCE 2006). Depending on the manufacturer, ORP meters need to be
cleaned often or equipped with automated cleaning (typically using air or water jets).
13.4 Online Instrumentation for Nutrient Control
Although it is more expensive and requires more attention than basic instrumentation, online
instrumentation specifically for nutrient control can be very important in optimizing nitrogen and
phosphorus removal processes. It can also lead to associated energy (and cost) savings via the use of
automated control systems. The next several sections discuss techniques and instrumentation available
for online monitoring for nutrient control
13.4.1 Nitrogen Compounds
Online monitoring of ammonia after the nitrification basin, in the effluent, and in return streams
from anaerobic digestion may be very useful for process control and optimization. Measurement of
nitrate and nitrite are important in determining 1) the load recycled to the anaerobic reactor of BPR
processes, 2) the load to the denitrification process, 3) the load in the denitrification process effluent
and 4) the control of external carbon addition for denitrification.
In general, manufacturers of online instruments for the various nitrogen species use one of
three methods (Palmer et al. 2007):
• Colorimetric approach. Colorimetric meters typically withdraw a sample from the wastewater
and give periodic readings and require the sample to be pretreated through a filter.
• Ion sensitive electrodes (ISEs). These are mounted directly in the water and can give real-time
readings. This method is similar to Standard Methods 4500-NH3-D (APHA et al. 2005)
Nutrient Control Design Manual 13-6 August 2010
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• UV absorbance or multiple wavelength UV absorbance spectrophotometers. This method is
based on Standard Methods nitrogen-persulfate digestion method 4500-N-B and nitrate,
ultraviolet spectrophotometric method 4500 NO3 - B (APHA et al. 2005).
Instruments based on colorimetry typically have an accuracy of 3 to 5 percent. ISEs are much faster and
are similarly reliable—within 2 to 5 percent. ISEs require more maintenance than colorimetric methods
because the electrode is placed directly in the wastewater stream. Electrodes can also experience
interference and have a limited pH range in which they will function.
The most common measurement technique for nitrate and nitrite is UV absorbance. Online
instruments direct UV light through the sample and measure the amount of light that is absorbed at a
specific wavelength. Because suspended solids interfere with the measurement method, filtration or
solids compensation is needed. The chemical oxygen demand (COD) of the water can also interfere with
spectrometer readings. To account for these interferences, manufacturers have incorporated multi-
wavelength measures or recommend membrane filtration (e.g., ultrafiltration) prior to sample analysis.
ISE and colorimetric techniques can also be used to measure nitrate and nitrite; however, the UV
absorbance method is most common (WEF and ASCE 2006).
Palmer et al. (2007) presents test data on commercially available in-situ UV absorption analyzers
for nitrate. The authors evaluated accuracy in measurements by comparing them to lab values. They
also collected information on capital costs, installation requirements, calibration requirements, and
annual maintenance costs. Some key results of the study are summarized in Table 13-2.
Table 13-2. Comparison of On-Line Nitrate Analyzers 1
Manufacturer
Endress+Hauser
HACH
HACH
s::can
Messtechnik
GmbH
Wedgewood
Analytical/E+H
Model
Stamo-Sens
CNM750/CNS70
Evita Insitu 5100
(Danfoss) 2
NITRATAXplussc
82 N Nitrolyser sensor
/ 8580 Constat
analyzer
STIP-scan
Mean Deviation from
Laboratory Conformance
Test
-0.08
2.88
0.68
0.26
0.55
Correlation to
Laboratory
Conformance Test
(R2)
0.92
0.02
0.97
0.91
0.61
Number of Field-
Test Maintenance
and Calibration
Events
None
3
1
3
3
1. Note that results based on test data from one facility
2. Hach Evita Insitu 5100 sensor is no longer on the market (discontinued in 2007)
Source: Palmer et al. 2007
13.4.2 Phosphate and Total Phosphorus
Monitoring phosphate (PO4) and total phosphorus (TP) is important in establishing the
phosphorus load on the plant, to control the phosphorus removal processes, and for compliance with
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permits. Measurements may be taken in the influent, effluent, anaerobic zone, anoxic zone, aerobic
zone, and return streams from digestion processes. Most online instruments measure PO4. Note that a
reading of 3 milligrams per liter (mg/L) as PO4 is roughly equivalent to 1 mg/L of phosphorus. Operators
should consider reporting results as P rather than as PO4, i.e as PO4-P. to avoid confusion (WEF and ASCE
2006).
Phosphate monitors typically use colorimetric techniques, especially for measurement at low
(0.1 mg/L or less) concentrations. The two most commonly used techniques are the Molybdenum blue
technique and the molybdovanadate technique, also called the "yellow" method. The molybdenum
blue technique generally has a range of 0.01 to 5 mg/L PO4-P, while the yellow method has a range of
0.1 to 50 mg/L PO4-P. Both monitors typically have accuracy and repeatability of approximately 2
percent of reading. The yellow method is more common due to the relative simplicity of the instrument
and reagents, but the blue method is required for confirming very low orthophosphate concentrations
in the plant effluent (WEF and ASCE 2006).
Total phosphorus instruments typically convert polyphosphate and organic phosphorus to
phosphate and then measure phosphate using one of the above mentioned methods. Because the
conversion steps require high temperatures and pressures, online total phosphorus monitors are very
expensive and typically only used for effluent monitoring, if at all. Many plants measure
orthophosphate in the effluent as a surrogate parameter. Accurate estimates of total phosphorus can
be obtained by combining it with the effluent TSS measurements.
Neupane et al. (2009) reports very recent findings in a study of orthophosphate and nitrogen
species analyzers as part of a pilot study of a Moving Bed Bioreactor (MBBR) at the Blue Plains Advanced
WWTP in Washington, DC. The ChemScan analyzer (UV 6101) was used to measure orthophosphate at
seven locations and compared to laboratory analysis of grab samples for a total of 6 months. The
ChemScan uses UV absorbance for all parameters and reports a detection limit of 0.05 mg/L for
orthophosphate. Neupane et al. (2009) noted that to maintain instrument accuracy, frequent cross
checks with the lab and recalibration were required. Weekly checking and replenishing of reagents was
also needed. However, authors report that the ChemScan OP readings were very accurate at
concentrations between 0.5 and 1.0 mg/L. The ChemScan could not read orthophosphate accurately at
the lower end of its calibration range.
13.4.3 NADH (active biomass)
A new technology developed in the Netherlands uses fluorescence to measure changes in
intracellular nicotinamide adenine dinucleotide (NADH) of microorganisms, which can provide
information on the status of biological wastewater treatment processes such as nitrification and
denitrification. The NADH sensor requires minimal maintenance and can provide real time information
for process control (Weerapperuma and de Silva 2004). NADH and DO monitoring can be used to
achieve simultaneous nitrification and denitrification in the same basin.
13.4.4 Respirometry
Respirometry involves measurement of the oxygen uptake rate (OUR) by a biological treatment
culture. In bench-scale respirometry experiments, a sample of mixed liquor, possibly amended with an
organic substrate or ammonia, is placed in a sealed vessel. The rate of oxygen consumption within the
Nutrient Control Design Manual 13-8 August 2010
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vessel is monitored over time. Alkalinity often needs to be added to nitrifying cultures to maintain a
neutral pH.
Online respirometers can be set up in several modes for process control. They all require a
biomass source representing the biomass in the reactor receiving influent wastewater. This can be
biomass from the process itself, the return activated sludge line, or from an off-line pilot reactor (Love
2002). The biomass, substrate if added, liquid phase, and oxygen are contained in a flow through
system or well mixed batch reactor. It is important that sufficient oxygen be present in the liquid or gas
phase to prevent oxygen limiting conditions (Love 2000). Oxygen is measured during the experiments to
determine oxygen uptake rates under normal operating conditions.
Online respirometry has several applications for activated sludge and nitrification treatment
processes. It can be used instead of online DO measurements to determine and control aeration
requirements and reduce electricity needs. One study in Chicago showed that this control is technically
feasible using a feed forward control strategy (Tata et al. 2000). It can also be useful as an upset early
warning device for the nitrification process. As explained in Chapter 4 of this manual, nitrifying bacteria
are vulnerable to inhibition from a number of different compounds. As part of a WERF study of upset
early warning devices, Love (2000) reviewed and documented several commercial devices that monitor
for inhibition of nitrification activities based on ammonia consumption or ratio of ammonia to BOD.
13.5 Types of Control
Regardless of the community they serve and treatment configuration, all plants experience
changes in external conditions, such as flow rate and constituent loading, and internal conditions, such
as growth of certain types of organisms. Control systems for wastewater treatment attempt to respond
to these changes by maintaining control parameters (e.g., DO) at target values by changing manipulated
parameters (e.g., aeration rate).
In the past and still the case for many (especially smaller) utilities, plants were operated
manually with operators collecting samples from various locations at the plant, bringing them back to
the laboratory for analysis (which could take several hours), evaluating the results, and changing
operating parameters to maintain desired control parameters. Automated systems essentially perform
these same tasks using online analyzers, computers, actuators, and communications systems.
Automated control can significantly reduce workload while at the same time improve control of process
parameters and thus, improve treatment plant performance, often at reduced operating costs.
There are several different types of control systems. Selection of the most appropriate one
depends on many factors including the specific control and manipulated parameters. The main types of
control—feed-forward, feedback, cascade, and advanced—are discussed below.
13.5.1 Feed-forward
Feed-forward control adjusts a parameter based on a measured change in another variable. The
parameter is adjusted in proportion to the measured variable. The value of the adjusted parameter can
be represented by:
Qa = K*Qm Eq. 13-1
Nutrient Control Design Manual 13-9 August 2010
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Where Q^ is the adjusted parameter, K is a constant, and Qm is the measured parameter. Feed-forward
control is commonly used when a parameter can be set based on flow rate. Examples include chemical
addition processes and return activated sludge flow rates.
13.5.2 Feedback
Feedback control is based on adjusting a parameter to correct an undesirable situation. The
adjusted parameter is changed in response to the measured parameter departing from a desired value
or set point. Dissolved oxygen control in aerated zones is often feedback type.
There are several types of feedback control including On/Off, proportional, integral,
proportional-integral, proportional-derivative, and proportional-integral-derivative. The control types
vary in how the adjusted parameter is changed in response to the departure of the measured parameter
from the setpoint.
On/Off control is the simplest type of feedback control. In it, a piece of machinery is turned on
when the measured variable falls outside of a set range and turned off when the measured variable
returns to that range. Examples include turning on air blowers when DO level in the aeration basin
drops below a set level. Proportional control adjusts a parameter in proportion to the departure of the
measured parameter from a setpoint. Integral control varies the adjusted parameter in proportion to
the area under a plot of the difference from the offset over time. Proportional integral control is a
combination of proportional and integral control. Proportional-derivative control adds a response that
is proportional to the derivative of the difference between the measured value and the setpoint. The
added derivative control makes the system much faster to respond, but also more likely to experience
chatter as the controller searches for the correct value around the setpoint. This type of control is not
as popular. Proportional integral-derivative control combines all three feedback control methods.
While this type of control combines the benefits of all three types of control, it can be harder to tune the
controller because of the larger number of variables.
13.5.3 Feed-forward and feedback
Feed-forward and feedback control combines the two control strategies. The controller initially
responds to a disturbance from the set point using a feed-forward control strategy. The response is
then fine-tuned using a feedback control. Again, while combining strengths of the two strategies, this
type of control can be difficult to tune.
13.5.4 Cascade
In cascade control, one feedback controller sets the setpoint for a second controller. This
configuration is used where two control loops may interact with each other. For example, control of air
controllers and control of DO concentrations interact and can be a candidate for this type of control.
13.5.5 Advanced Control
Advanced control systems apply unique algorithms specific to a given treatment configuration
through use of a computer or other programmable devices. They use input from many sensors and
Nutrient Control Design Manual 13-10 August 2010
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mathematical models, equations, or condition matrices to manipulate equipment to obtain the desired
outcome. Other straightforward control systems such as feed-forward or feedback are often part of an
advanced system. Control can be based on mathematical models such as ASM1 (see Chapter 10 for
additional information on the many uses of models for plant design, operation, and optimization).
13.6 Control Equipment—SCADA
Supervisory Control and Data Acquisition (SCADA) systems involve a central computer that is
linked to various devices by means of a communications system. The central computer receives and
sends information by means of the communications system, which can be through phone lines or radio
waves. SCADA systems can be used for continuous process control, data acquisition, supervisory
control, alarming, historical data collection, as an information database, and to provide security for
stored information.
A SCADA system will constantly monitor and may be used to control the wastewater process.
The central server can communicate with individual sensors and make control decisions based on
process logic contained in the server. Communication with operator consoles occurs through either
proprietary protocols, TCP/IP protocols, or through OPC protocols. Program logic should meet IEC
61131-3 standards.
The SCADA system will also monitor all sensors throughout the plant. The data are collected
and can then be used for other functions such as control, display for operators, or historical storage.
Supervisory control is one of the essential features of a SCADA system. It allows operators to
make changes to the automatic control of the plant by changing set points and starting and stopping
machinery. The SCADA system will also allow the operator to monitor the various processes from a
central point.
SCADA systems also perform the important function of notification of problems through a
hierarchy of alarms. The SCADA system will record alarm conditions, prioritize alarms, notify operators
through audible and visible alarms, and allow alarm acknowledgement. SCADA systems can also be set
to notify off-site personnel through phone messages or text messages.
SCADA systems can archive all data collected such as process variables, alarm logs, and
operating conditions. These data can be displayed at any time for troubleshooting and analysis of the
plant. Data can be analyzed using functions on the SCADA computer and can also be used as inputs for
wastewater models and other analytic tools. The SCADA system stores all information in a central
database where it can be accessed by all plant personnel for various purposes including preparing
reports, troubleshooting, and monitoring the plant.
Security is an essential component of a SCADA system. The security is set up to prevent
unauthorized use and also to specify what function each authorized user can perform depending on that
person's responsibilities.
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13.7 References
Akin, B.S. and A. Ugurlu. 2005. Monitoring and Control of Biological Nutrient Removal in a Sequencing
Batch Reactor. Process Biochemistry. 40(8): 2873-2888.
American Public Health Association (APHA), American Water Works Association (AWWA) & Water
Environment Federation (WEF). 2005. Standard Methods for the Examination of Water and
Wastewater, 21st Edition. Eds. A.D. Eaton, L.S. Clesceri, E.W. Rice, A.E. Greenberg, and M.H. Franson.
Baeza, J.A., D. Gabriel, and J. Lafuente. 2004. Effect of Internal Recycle on the Nitrogen Removal
Efficiency of an Anaerobic/Anoxic/Oxic (A2O) Wastewater Treatment Plant. Process Biochemistry.
39(11): 1615-1624.
Baeza, J.A., E.G. Ferreira, and J. Lafuente. 2003. Knowledge Based Supervision and Control of Wastewater
Treatment Plant Real Time Implementation. Water Science and Technology. 14(12): 129-137.
Barnard, J. L, A. Shaw, J.B. Watts. 2003. Optimized Control of Nitrifying Suspended Growth Systems
Using Respirometry. Proceedings of the Ozwater Convention and Exhibition, AWA 20th Convention,
Perth, Australia. Sydney: Australian Water Association.
Ekster, A. and I. Rodrfguez-Roda. 2003. The Effect of Sludge Age on the Operation Cost of Activated
Sludge System. Presented at WEFTEC 2003.
Hill, R.D., R.C. Manross, E.V. Davidson, T.M. Palmer, M.C. Ross, S.G. Nutt. 2002. Sensing and Control
Systems: A Review of Municipal and Industrial Experiences. WERF Report 99-WWF-4. WERF.
Alexandria, VA.
Isaacs, S., M. Henze, and M. Kummel. 1995. An Adaptive Algorithm for External Carbon Addition to an
Alternating Activated Sludge Process for Nutrient Removal from Wastewater. Chemical Engineering
Science. 50(4): 617-629.
Kestel, S., G.J. Duffy, M. Gray, T. Stahl, and G. Lee. 2009. DO Control Based on On-line Ammonia
Measurement. Presented at Nutrient Removal 2009. Washington, DC. WEF.
Lara, J.M.V. and B.E.A. Milani. 2005. Control of Nitrate Flow in Pre-nitrification Systems Using Long Range
Identification for Predictive Control. Decision and Control. 2005 European Conference.
Love, N.G. 2000. A review and Needs Survey of Upset Early Warning Devices. WERF Report 99-WWF-2.
WERF. Alexandria, VA
Liu, W., G. Lee, and J. Goodley. 2003. Using On-Line Ammonia and Nitrate Instruments to Control
Modified Ludzack-Ettinger (MLE) Process. Presented at WEFTEC 2003.
Ma, Y., Y. Peng, and S. Wang. 2005. New Automatic Control Strategy for Sludge Recycling and Wastage
for the Optimum Operation of Pre-denitrification Processes. Journal of Chemical Technology and
Biotechnology. 81(1): 41-47.
Nutrient Control Design Manual 13-12 August 2010
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Manning, A. 2008. Automation, Instrumentation Generations 2, 3 and Beyond: Let's Get it Right This
Time. Presented at WEFTEC 2008.
Meyer, J. 2008. Fiesta Village Advanced Wastewater Treatment Plant: Manager's Perspective on
Following Separate Stage Denitrification with Fixed Film Systems for Low Nitrogen. Presented at
WEF/WERF: Demonstrated Processes for Limit of Technology Nutrient Removal: Achievable Limits and
Statistical Reliability. October 18, 2008. Chicago, IL.
Neupane, D., M. Peric, B. Stinson, M. Laquidara, S. Kharkar, N. Passarelli, S. Murthy, W. Bailey, R.
DerMinassian, and M. Sultan. 2009. Comparative Study of On-line Instruments for Monitoring Low
Nutrient Concentrations. Presented at Nutrient Removal 2009. Washington, DC. WEF.
Palmer, T. 2007. On-Line Nitrogen Monitoring and Control Strategies. Report: 03-CTS-8. Alexandria, VA:
WERF.
Puznavan, N., S. Zeghel, and E. Reddet. 1998. Simple Control Strategies of Methanol Dosing for Post-
denitrification. Water Science and Technology. 38(3): 291-297.
Samuelesson, P., Carlsson, B. 2001. Control of Aeration Volume in an Activated Sludge for Nutrient
Removal. Proceedings of the first IWA Conference on Instrumentation, Control and Automation.
Malmo, Sweden, June 3-7. pg. 429 - 437
Tata, P., K. Patel, S. Soszynski, C. Lue-Hing, K. Cams, and D. Perkins. Potential for the Use of On-line
Respirometry for the Control of Aeration. Presented at WEFTEC 2000.
Tchobanoglous, G., F. L. Burton, and H.D. Stensel. 2003. Wastewater Engineering: Treatment and Reuse.
New York, NY: McGraw-Hill.
Weerapperuma, D. and V. de Silva. 2004. On-line Analyzer Applications for BNR Control. Presented at
WEFTEC 2004.
WEF, 2007. Operation of Municipal Wastewater Treatment Plants - MOP 11, 6th Ed. Water Environment
Federation. Alexandria, VA. WEFPress.
WEFandASCE. 1998. Design of Municipal Wastewater Treatment Plants - MOP 8, 4th Ed. Water
Environment Federation and American Society of Civil Engineers. Alexandria, VA: WEF.
WEF and ASCE. 2006. Biological Nutrient Removal (BNR) Operation in Wastewater Treatment Plants -
MOP 29. Water Environment Federation and the American Society of Civil Engineers. Alexandria, VA:
WEFPress.
WEF and ASCE. 200. Design of Municipal Wastewater Treatment Plants - WEF Manual of Practice 8 and
ASCE Manuals and Reports on Engineering Practice No. 76, 5th Ed. Water Environment Federation,
Alexandria, VA, and American Society of Civil Engineers Environment & Water Resources Institute,
Reston, Va.
Yuan, Z., H. Bogaert, P. Vanrolleghem, C. Thoeye, G. Vansteenkiste, and W. Verstaete. 1997. Journal
Environmental Engineering. 123(11): 1080-1086.
Nutrient Control Design Manual 13-13 August 2010
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Yuan, Z. and J. Keller. 2003. Integrated Control of Nitrate Recirculation and External Carbon Addition in a
Predenitrification System. Water Science and Technology. 48(11-12): 345-354.
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14. Sustainable Nutrient Recovery and Reuse
Chapter 14 covers:
14.1 Introduction
14.2 Separating and Treating Waste On-Site
14.3 Using Wastewater Treatment Byproducts
14.4 References
14.1 Introduction
Sustainable nutrient recovery and reuse is gaining national and international attention as
wastewater utilities look for ways to decrease energy costs and greenhouse gas emissions, utilize excess
capacity, generate new revenue, and address ever more stringent regulatory requirements. This
evolution in thinking is moving wastewater treatment to enhanced energy efficiency and changing the
role of wastewater treatment facilities from waste generators to resource providers.
In the United States and abroad, research on innovative sustainable nutrient recovery systems
is changing the way liquid and solid waste is being treated and reused. This chapter will describe some
of the on-site and centralized technologies and approaches currently being investigated and
implemented.
14.2 Separating and Treating Waste On-Site
A number of European countries including Switzerland, Sweden, and the Netherlands are
conducting research on innovative sustainable nutrient recovery systems. The concept behind these
new technologies is to separate toilet waste before it leaves the home or building and mixes with the
larger waste stream to be carried to wastewater treatment plants (WWTPs). Recent studies have shown
that about 80 percent of the nitrogen and 50 percent of the phosphorus in wastewater are derived from
urine, even though urine makes up only 1 percent of the volume of wastewater (Larsen and Leinert
2007). Separating the urine from wastewater could offer various advantages: WWTPs could be built on
a smaller scale, water bodies would be better protected from nitrogen and phosphorus pollution,
nutrients could be recycled for agricultural use, and various constituents of concern including hormones
and pharmaceutical compounds found in urine could be removed without being mixed with wastewater
and released to the environment. A major benefit would be reduced energy consumption at WWTPs as
a result of reduced treatment requirements for nitrogen. Also, separating 50 to 60 percent of urine
could reduce in-plant nitrogen gas discharges and result in fewer impurities in methane captured from
sludge digestion.
Organizations such as the Swiss Federal Institute of Aquatic Science and Technology (EAWAG)
are currently experimenting with the development and application of "NoMix technology" to separate
urine from solid waste at the toilet bowl. While similar in size and shape to current toilets, this new
technology has two waste pipes—a small front one that collects and diverts urine into a storage tank,
and a larger rear waste pipe that operates like a standard toilet. The first of these toilets were installed
in two "eco-villages" in Sweden in 1994 and since then have spread to other locations throughout the
country and to Denmark, the Netherlands, and Switzerland. The concept is now taking hold in Austria
Nutrient Control Design Manual 14-1 August 2010
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and Germany. While the urine, or "urevit," can be spray-applied directly onto agricultural fields, a
company called Grontmij in the Netherlands transports stored urine to a special treatment plant where
the phosphate is precipitated out as struvite (magnesium-ammonium-phosphate) and used as a
fertilizer.
EAWAG is also experimenting with a pilot decentralized basement sewage plant in which
domestic wastewater is treated in a membrane bioreactor (MBR) so it can be reused for flushing the
toilets or watering the garden and the sludge is composted. While still experimental, some of these
technologies may have practical future applications if widely applicable low-cost solutions can be found
for urine transport, or stable and cost-effective technologies can be developed for decentralized
treatment. While studies of consumer attitudes and acceptance appear to be positive, technological
improvements are still needed to prevent clogging in pipes, to identify best treatment options that can
be applied in practice, and to identify how and where to convert urine to fertilizer.
14.3 Using Wastewater Treatment Byproducts
Sustainability concerns are also driving the wastewater treatment industry to look at liquids and
sludge as renewable resources. Historically, agricultural use has been the traditional approach for
disposal of municipal sludge due to its high nutrient content for fertilizing crops and its low cost. As
scientific advances detect smaller and smaller quantities of chemicals, heavy metals, pathogenic
microorganisms, Pharmaceuticals, and personal care products, concerns are being raised about
continuing this practice.
Researchers are exploring alternative approaches to treating and using the valuable products
generated from sewage treatment byproducts such as energy extracted from anaerobic digestion,
construction materials such as cement and bricks, and nutrients such as phosphorus that can be
extracted from sludge and used as fertilizer. Liquids generated through the treatment process are also
being harvested for nutrients for fertilizers and reused as non-potable sources of water for agriculture,
landscape irrigation, and indirect potable reuse such as aquifer storage and recovery. It has also been
demonstrated that struvite will form in anaerobic digesters treating biological phosphorus removal
(BPR) sludges and will remain in the sludge rather than subsequently precipitate in the outflow pipes.
The sludge can then be sterilized by further processing such as composting. Additionally, several
processes have been developed for the precipitation of struvite from highly concentrated waste streams
such as digester supernatants and dewatering filtrates. Several different processes have been proposed
that rely on precipitation of the phosphorus as either struvite or calcium phosphate. Scope (2004)
reported that at the time, work was underway in Italy, Germany, the Netherlands, and Canada.
In February 2008, the non-profit Global Water Research Coalition, an international water
research alliance formed by 12 world-leading research organizations, released a report in collaboration
with USEPA and the Water Environment Research Foundation (WERF) titled, State of Science Report:
Energy and Resource Recovery from Sludge (Kalogo and Monteith 2008). The report focuses on:
• The international situation of energy and resource recovery from sludge.
• How the use of different sludge treatment processes affects the possibility of recovering
energy and/or materials from the residual sludge.
• The influence of market and regulatory drivers on the fate of the sludge end-product.
Nutrient Control Design Manual 14-2 August 2010
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• The feasibility of energy and resource recovery from sludge.
• The social, economic, and environmental performance (triple bottom line orTBL
assessment) of current alternative technologies.
In the report, energy recovery technologies are classified into the following processes: sludge-
to-biogas, sludge-to-syngas, sludge-to-oil, and sludge-to-liquid. The technologies available for resource
recovery discussed in the report include those to recover phosphorus, building materials, nitrogen, and
volatile acids. The report, which covers both established and emerging technologies, forms the basis of
the coalition's strategic research plan on energy and recovery from sludge. As a technical resource, it
provides a valuable overview of sludge disposal practices in various countries such as the United States,
the Netherlands, the United Kingdom, Germany, Sweden, Japan, and China, as well as a number of
treatment processes for resource recovery.
Examples of these processes are described below.
14.3.1 Durham, OR, Advanced Wastewater Treatment Facility
Operated by Clean Water Services, the Durham Advanced Wastewater Treatment Facility in
Tigard, OR, has incorporated a new technology from Ostara Nutrient Recovery Technologies, Inc. to
extract phosphorus and other nutrients from wastewater and recycle them into an environmentally safe
commercial fertilizer called Crystal Green®. Rather than reprocessing the sludge liquids back through
the wastewater system, the Ostara proprietary fluidized bed reactor uses magnesium to recover
ammonia and phosphate and turns them into a fertilizer pellet. By extracting these nutrients, the
concrete-like scale of pure mineral called "struvite" (magnesium-ammonium-phosphate) no longer clogs
the facility's pipes, valves, and other equipment, thereby increasing plant capacity and reducing
maintenance costs, while creating a revenue-generating by-product. Struvite is a very valuable slow-
release fertilizer. Full-scale plant operation began in May, 2009. It is projected that more than 90
percent of the phosphorus in the return streams will be removed and more than 500 tons of struvite
fertilizer will be produced annually and sold to commercial enterprises. Clean Water Services expects
that the initial investment of $2.5 million will be paid back within five years. For more information, see
the Clean Water Services website at
http://www.cleanwaterservices.org/AboutUs/WastewaterAndStormwater/Ostara.aspx
In addition to phosphorus recovery, the Ostara process:
• Reduces pollution
• Reduces chemical use
• Removes about 20 percent of the nitrogen in the return streams
• Generates revenue from the sale of the fertilizer
• Increases the reliability and capacity of the wastewater facility
• Reduces operation and maintenance costs
• Reduces greenhouse gas emissions through reduced energy use
• Potentially generates carbon credits
Nutrient Control Design Manual 14-3 August 2010
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The process does not have a significant impact on how wastewater is treated. Moreover, it does not
require further chemical phosphorus precipitation from the return stream while it removes some of the
ammonia that will otherwise need to be oxidized and denitrified in the main stream. This technology is
particularly suited for WWTPs that employ biological phosphorus removal and anaerobic sludge
digestion.
14.3.2 East Bay Municipal Utility District, CA
East Bay Municipal Utility District (EBMUD) has applied its biosolids to agricultural lands for
decades, returning nutrients to the soil. Biosolids are rich in nutrients such as nitrogen and phosphorus,
and the benefits of land application are well known. However, many agricultural communities are
adopting more stringent land application standards because of concerns regarding pathogens. EBMUD
has conducted research to develop economical alternatives for the production of Class A ("pathogen
free") biosolids, including thermophilic anaerobic digestion. EBMUD holds a patent for a thermophilic
anaerobic digestion process. In 2005, EBMUD developed an environmental management system (EMS)
to identify cost and environmental savings, ensure regulatory compliance, and increase public
awareness and involvement in the biosolids program. As part of this program EBMUD undergoes third
party audits. For results, see EBMUD's website at http://www.ebmud.com/our-water/wastewater-
treatment/wastewater-treatment-programs/environmental-management-svstem.
EBMUD also captures and reuses biogas from its anaerobic digesters. Methane from the biogas
is used to fuel a 6 megawatt on-site power plant. To increase methane production, EBMUD
implemented a trucked waste program whereby they accept liquid and solid waste streams from
outside their service area and process it at the plant to increase methane production. Wallis et al.
(2008) reported a doubling of energy production over the past six years. EBMUD plans to add two 4.5
megawatt turbines that will allow them to generate 100 percent of their energy needs onsite by the end
of 2010. For more information on EBMUD's sustainability programs, see
http://www.ebmud.com/environment/sustainability.
14.4 References
Kalogo, Y., and H. Monteith. 2008. State of Science Report: Energy and Resource Recovery from Sludge.
Prepared for Global Water Research Coalition, by WERF, STOWA, and UK Water Industry Research
Limited.
Larsen, T.A., and J. Leinert, Eds. 2007. Novaquatis Final Report. NoMix-A New Approach to Urban
Water Management. Switzerland: EAWAG, Novaquatis.
SCOPE. 2004. Newsletter No. 57. July. Centre Europeen d'Etudes sur les Polyphosphates. Brussels,
Belgium. Available online: http://www.ceep-
phosphates.org/Files/Newsletter/Scope%20Newsletter%2057%20Struvite%20conference.pdf
Wallis, M., M. Ambrose, and C. Chan. 2008. Climate Change: Charting a Water Course in an Uncertain
Future. Journal AWWA 100:6, June 2008.
Water & Wastewater News. 2008. Oregon Plant Recycles Nutrients into Fertilizer. L.K. Williams, Ed.
October 2, 2008. Available online: http://www.wwn-online.com/print.aspx?aid=68074.
Nutrient Control Design Manual 14-4 August 2010
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Appendix A.
Recommendations for Methanol Safety
-------
Appendix A. Recommendations for Methanol Safety
Methanol (CH3OH) is a relatively simple compound with a single carbon bond. Clear and
colorless at room temperature, it is fully soluble in water and has a faintly sweet pungent odor. It is
typically made from natural gas, although it can be generated using renewable materials such as wood.
As noted in Chapter 4, methanol is the most common external carbon source for denitrification
because of its historically low cost and relatively low cell yield. The use of methanol, however, poses
several safety and health risks:
• Risk of Fire. Methanol is highly flammable, with a flashpoint of 12 °C (54 °F). Even mixtures
containing as little as 25 percent of methanol by weight are flammable. Because methanol is
heavier than air, vapors can travel to a source of ignition and flash back. Vapors can also
accumulate in confined spaces and may explode if ignited. Containers of methanol solution can
rupture if exposed to fire or excessive heat for a long duration. Methanol burns with a clear blue
flame. During daylight hours, methanol fires may not be visible to the naked eye. The U.S.
Department of Transportation (DOT) ranks methanol and solutions of methanol with strengths
of greater than 25% as Class 3 Flammable Liquids, meaning that they could ignite under almost
all ambient conditions.
• Health Risks: Methanol is irritating to the eyes, skin, and respiratory track and is very harmful if
swallowed or absorbed through the skin. Swallowing as little as 50 milliliters of methanol (1/4
of a cup) can be fatal. Inhalation is the most common form of exposure and can lead to
respiratory problems. Methanol can also affect the nervous system and cause a loss of
consciousness. Repeated exposures to low concentrations may accumulate in the body and
cause illness. For additional information on the health effects of methanol, see information
from EPA's Integrated Risk Information System (IRIS) at
http://www.epa.gov/iris/subst/0305.htm.
This appendix contains recommendations for methanol handling. Literature reviewed for the
appendix included a report by the U.S. Chemical Safety and Hazard Investigation Board on the Bethune
Point Wastewater Treatment Plant accident, a Water Environment Federation (WEF) webcast entitled,
Methanol Safety: Understanding and Managing Risks, and proceedings from WEFTEC2008. When
designing and operating a methanol feed system, plants should always follow guidelines in the
manufacturer's Material Safety Data Sheet (MSDS). A good clearinghouse of information on methanol
is provided online by the Methanol Institute, http://www.methanol.org/. Wastewater utilities should
also always follow Occupational Safety & Health Administration (OSHA) regulations and National Fire
Protection Association (NFPA) codes and standards.
A.I Preventative Measures
It is imperative to keep open flames, sparks, and oxidants away from methanol. The best way to
prevent fire is to eliminate the source of ignition. Operators and maintenance personnel should always
use hand tools that do not spark.
Nutrient Control Design Manual A-2 August 2010
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Methanol storage and handling areas should be equipped with the following:
• Local exhaust
• Safety shower and eyewash station
• Smoke and/or heat detectors
• Fire suppression system (extinguishing materials including dry chemical powder, alcohol
resistant foam, and water)
• Supply of absorbent, non-combustible materials in event of small spill
• Control measures to limit spread of spill or fire (i.e., secondary containment)
Wastewater utilities should consider monitoring for methanol vapors using detector tubes, electronic
instruments, absorbent tubes, or biological monitoring (Dolan 2007).
Methanol is non-corrosive to most metals, the exceptions being lead, magnesium, and platinum.
Mild steel is commonly used for methanol storage. Vapor controls include internal or external floating
roofs, inert gas blanketing (such as dry nitrogen), and pressure vacuum relief to control tank pressures
(Methanex 2006). Flame arresters can be attached to vent pipes to allow passage of methanol vapors
but prevent flame from moving into the pipe by transferring heat to metal wires or plates. Flame
arresters may become ineffective if they contain any blockages or corrosion.
When designing and constructing a new methanol feed system, consider involving local fire
officials in the process. At the Stamford Water Pollution Control Authority (WPCA), the Fire Marshall
was present at the start up of the methanol system and witnessed the initial foam fire suppression
system test. The Fire Marshall's office periodically inspects the facility and alarm, sprinkler, and foam
suppression systems (Brown 2007).
Operations manuals should specify regular inspections and maintenance of methanol handling
equipment. Operators should undergo regular training on methanol handling and safety.
A.2 Addressing a Spill
The method for cleaning up a methanol spill depends on its size. For small spills (generally less
than the contents of a 55 gallon drum), methanol can be absorbed with earth, sand, or other non-
combustible material. Material should be stored in a sealed container and disposed of at a facility
licensed to handle hazardous waste. A fine water spray or vapor suppressing foam can be used to
reduce vapor. Personnel should wear protective gloves, fire retardant clothing (e.g., rubber), and safety
goggles or other form of eye protection.
For larger spills, it is important to follow local emergency protocols and treat the spill as a
hazardous situation. Wastewater utility personnel should eliminate ignition sources, secure the area up
to a 500 meter radius, restrict access, move personnel upwind, and decontaminate personnel and
equipment. All personnel allowed entry should wear a full chemical suit, compressed air breathing
apparatus (CABA), rubber boots, and gloves. Although methanol biodegrades quickly in soil and water,
measures should be taken to prevent methanol from entering waterways, sewers, basements, or
confined areas.
In the event of a fire, utility personnel should follow written emergency response procedures
and should immediately notify local authorities including the fire department and Emergency
Nutrient Control Design Manual A-3 August 2010
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Management Services (EMS) personnel. Powder extinguisher, alcohol-resistant foam, carbon dioxide,
and large amounts of water can be used to suppress the fire in the event of an emergency. Water can
be less effective than other methods; however, it may be useful in diluting methanol to a non-
flammable level. Avoid using straight streams of water to prevent spread of contamination. Drums
should be kept cool by spraying them with water to prevent explosion.
A.3 References
Brown, J.A. 2007. Methanol Storage and Delivery System. WERF External Carbon Workshop. Stamford
WPCA. December 12, 2007. Washington, DC.
deBarbadillo, C., J. Barnard, S. Tarallo, and M. Steichen. 2008a. Got Carbon? : Widespread Biological
Nutrient Removal is Increasing the Demand for Supplemental Sources. Water Environment &
Technology. Alexandria, VA: WEF. 20(1): 48-53.
deBarbadillo, C., P. Miller, S. Ledwell. 2008b. A Comparison of Operating Issues and Dosing
Requirements for Alternative Carbon Sources in Denitrification Filters. In Proceedings of WEFTEC
2008. Alexandria, VA: WEF.
Dolan, G. 2007. Methanol Safe Handling. WERF Carbon Sources Seminar. Methanol Institute.
December 12, 2007. Washington, DC.
EDIE. Methanol Storage and Dosing for Denitrification. Gee& Company Product Information.
Environmental Data Interactive Exchange. Surrey, England: Faversham House Group. Accessed July
24,2008. Available online: http://www.edie.net/products/view entry.asp?id=2401
IPCS. 2004. Methanol Data Sheet: 0057. International Programme on Chemical Safety. Available
online:
http://www.ilo.org/public/english/protection/safework/cis/products/icsc/dtasht/jcscOO/icsc0057.ht
m
Methanex Corporation. 2006. Technical Information & Safe Handling Guide for Methanol, Version 3.0.
Vancouver, British Columbia, Canada: Methanex.
Methanex Corporation. Methanol Safe Handling and Storage, Distributed Generation Code Workshop.
Vancouver, British Columbia, Canada: Methanex.
Methanol Institute. Methanol Emergency Response Factsheet. Arlington, VA: Methanol Institute.
Methanol Institute. Methanol Health Effects Factsheet. Arlington, VA: Methanol Institute.
Methanol Institute. 2005. Methanol and Wastewater Denitrification. Arlington, VA: Methanol
Institute.
Methanol Institute. Wastewater Treatment with Methanol Denitrification. Arlington, VA: Methanol
Institute.
Nutrient Control Design Manual A-4 August 2010
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NPI. 2007. Methanol: Rank 24 of 90 Substances. Australian Government. Department of the
Environment and Water Resources. National Pollutant Inventory. Last updated October 26, 2007.
Available online: http://www.npi.gov.au/database/substance-info/profiles/54.html
Onnis-Hayden, A. and A.Z. Gu. 2008. Comparisons of Organic Sources for Denitrification:
Biodegradability, Denitrification Rates, Kinetic Constants and Practical Implication for Their
Applications in WWTPs. In Proceedings of WEFTEC2008. Alexandria, VA: WEF.
Terra Nitrogen Corporation. 2001. Material Safety Data Sheet: Methanol. MSDS Number 2016. Sioux
City, Iowa: Terra.
U.S. Chemical Safety and Hazard Investigation Board. 2003. Public Board Meeting. March 6, Houston,
Texas.
U.S. Chemical Safety and Hazard Investigation Board. 2007. Investigation Report: Methanol Tank
Explosion and Fire, Bethune Point Wastewater Treatment Plant, City of Daytona Beach, FL, January 11,
2006. Report 2006-03-l-FL
WEF. 2008. Methanol Safety: Understanding and Managing Risks. Webcast organized by the WEF
Safety, Security, and Occupational Health Committee (SSoHC) with the WEF Municipal Wastewater
Design Committee, the Methanol Institute, and the Chemical Safety Board. 23 July 2008.
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Appendix B.
Organic Compounds and Inhibitory Concentrations
to Nitrification
-------
Appendix B. Organic Compounds and Inhibitory Concentrations to Nitrification
Organic Compound
Acetone
Allyl alcohol
Allyl chloride
Allyl
Allyl thiourea
AM (2-amino-4-
Amino acids
Aminoethanol
Aminoguanidine
2-Aminophenol
4-Aminophinol
Aminopropiophen
Aminotriazole
Ammonium
Aniline
1-Arginine
Benzene
Benzidine
Benzocaine
Benzothiazzole
Benzylamine
Benzyldimethyldod
Benzylthiuronium
2.2' Bipyridine
Bisphenol A
Bromodichloropro
2-Bromophenol
4-Bromophenol
n-Butanol
Cadmium
Carbamate
Inhibitory Concentration (mg/L)
2000
19.5
180
1.9
1.2
50
1-1 000
12.2
74.0
0.27
0.07
43
70.0
1000
7.7
1.7
13.00
50.0
100
38.0
100
2.0
40.0
10.0
100
84.0
0.35
0.83
8200
14.3
2
Nutrient Control Design Manual
B-2
August 2010
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Organic Compound
Carbon disulphide
Chlorine
Chlorobenzene
Chloroform
2-Chloronaphthol
2-Chlorophenol
3-Chlorophenol
4-Chlorophenol
5-Chloro 1-
2-Chloro-6-
m-Cresol
o-Cresol
p-Cresol
Cyanide
Cyclohexylamine
Di-allyl ether
1,2-Dibromoethane
Dibromethane
1,2-
1,1-Dichloroethane
2,4-Dichloroethane
1,5-
2,3-Dichlorophenol
2,3-Dichlorophenol
2,6-Dichlorophenol
3,5-Dichlorophenol
1,3-
1,3-
Dicyandiamide
Dicyclohexylcarbod
Diethyl
Diguanide
Dimethylgloxime
Inhibitory Concentration (mg/L)
35.0
1
0.71, 500
18.0
14.3
2.70
0.20
0.73
0.59
11.0
01. -100
11.4
12.8
16.5
0.500
100
50.0
60
100
0.91
0.79
13.00
0.42
0.61
8.10
3.00
0.67
0.48
250
10.0
0.1
50.0
140
Nutrient Control Design Manual
B-3
August 2010
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Organic Compound
Dimethylhydrazine
Dimethyl p-
Dimethyl p-
2,4-Dinitrophenol
Diphenylthiocarbaz
Dithio-oxamide
Dodecylamine
Erythromycin
Ethanol
Ethanolamine
Ethyl acetate
Ethylenediamine
Ethyl urethane
Ethyl xanthate
Flavonoids
Guanidine
Hexamethylene
Histidine
Hydrazine
Hydrazine sulphate
Hydrogen sulfide
8-Hydroxyquiniline
Lauryl
1-Lysine
Mercaptobenzothi
Methanol
Methionine
n-Methylaniline
Methylhydrazine
Methyl
Methyl mercaptan
Methyl pyridines
2-Methylpyridine
Inhibitory Concentration (mg/L)
19.2
19.0
30
37.0
7.5
1
<1
50.0
2400
100
18
100
1000
10
0.01
4.7
85
5
58.0
200
50
1
118
4.0
3
160
9.0
<1
12.3
0.800
300
100
100
Nutrient Control Design Manual
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August 2010
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Organic Compound
4-Methylpyridine
Methylthiourea
Methyl thiuronium
Methylamine
Methylene blue
Monethanolamine
N- serve
Napthylethylenedi
Ninhydrin
p-Nitroaniline
P-
Nitrobenzene
4-Nitrophenol
2-Nitrophenol
2-Nitrophenol
Nitrourea
Panthothenic acid
Pentachloroethane
Perchloroethylene
Phenolics
Phenolic acids
p-Phenylazoaniline
Potassium
Potassium chlorate
Potassium
Potassium
n-Propanol
Purines
Pyridine
Primidines
Pyruvate
Resurcinol
Skatole
Inhibitory Concentration (mg/L)
100
0.455
1
100
30
>200
10
23
10.0
10.0
50.0
50.0
2.60
11.00
50.0
1.0
50
7.90
5.6
100
0.01
100
800
2500
6.0
300
20.0
50
10.0
50
400
7.80
7.0
Nutrient Control Design Manual
B-5
August 2010
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Organic Compound
Sodium azide
Sodium azide
Sodium arsenite
Sodium chloride
Sodium cyanate
Sodium cyanide
Sodium dimethyl
Sodium methyl
Sodium pluoride
Sodium
ST(sulfathiazole)
Strychnine
Sulphides
Tannin
Tetrabromobisphen
1,2,3,4-
1,2,4,5-
1,1,1,2-
1,1,2,2-
12,3,5,6-
Tetramethylammon
Tetramethyl
Thiamine
Thioacetamide
Thiocyanates
Thiosemicarbazide
Thiourea
Thiourea
1-Threonine
Threonine
2,4,6-
2,4,6-
2,4,6-
Inhibitory Concentration (mg/L)
23.0
20
2000
35000
100
1
13.6
0.90
1218
1
50
100
5.0
0.01
100
20.00
9.80
8.70
1.40
1.30
2200
5
0.530
500
0.180
0.760
1
3.6
5
50.0
7.70
50
2.5
Nutrient Control Design Manual
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August 2010
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Organic Compound
2,2,2-
1,1,2-
Trichloroethylene
Trichlorophenol
2,3,5-
2,3,6-
2,4,6-
Triethylamine
Trimethylamine
2,4,6-
1-Valine
Vitamins riboflavin,
Inhibitory Concentration (mg/L)
2.00
1.90
0.81
100
3.90
0.42
7.90
100
118
30.0
1.8
50
Source: Reprinted with permission from Biological Nutrient Removal Operation in Wastewater Treatment Plants, Copyright ©
2005, Water Environment Federation Alexandria, VA, www.wef.org. From Blum and Speece, 1991; Christensen and Harremoes,
1977; Hockenbury and Grady, 1977; Painter, 1970; Payne, 1973; Richardson, 1985; Sharma and Ahlert, 1977.
Nutrient Control Design Manual
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Nutrient Control Design Manual B-8 /August 2010
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Appendix C.
Mathematical Models for Wastewater Treatment
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Appendix C. Mathematical Models for Wastewater Treatment
Models are sets of equations which describe a physical system. Wastewater modeling has
advanced from a rudimentary understanding of the activated sludge process to complex systems of
equations representing all the processes occurring in a wastewater plant. Figure C-l shows a brief
outline of the development of modeling of activated sludge processes. Activated sludge modeling first
began to find widespread use in the wastewater industry after the International Water Association
(IWA) produced the first Activated Sludge Model (ASM1) in 1986. Since then many enhancements have
been added to ASM1 resulting in several improved versions of the basic ASM models. Several other
models have also been developed to address specific processes within the activated sludge process as
well as models for fixed film processes and settling. Powerful simulation packages have also been
developed, capable of linking several models representing each process in the plant from primary
settling through the activated sludge process to secondary clarification and sludge digestion.
Nutrient Control Design Manual C-2 August 2010
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Activated Sludge Modeling Timeline
Penfold ar-d
Morris [1912J
Bacteria
Generation
Time
Garrett and
Sowy&f |1952[
Mixed Culture
Kinetics
Busby ancJ
Ardrev.^ .|l?7Sl
First" Computer
Model (CSAAP)
Grady e~
al. (T986)
ASM1
henze et
al. (1999)
ASAA2d
Gujef et
al. (1999)
ASMS
I
19 1C
I
192C
I
193C
Ar
-------
Models can be either black box models or white box models. Black box models fit a set of
experimental data to an empirical equation. White box models are based on actual knowledge of the
processes and their description in mathematical terms. Although some work has been done using black
box models in wastewater, most commonly used models for design are white box models. This
appendix will focus on those white box models commonly used for design of wastewater nutrient
removal.
There are many excellent references on mathematical modeling of wastewater treatment
plants. Gernaey et al., (2004) provide a concise summary of existing models and their underlying
assumptions. Comeau and Takacs (2008) propose schematic representations for the mathematical
models and a standardized list of symbols to further understanding of the models' mechanisms, use, and
limitations. Additional information on fixed film mathematical models can be found in Wanner et al.,
(2006). References for individual models can be found in tables C-l and C-2.
C.I Description of Available Models
Settling
Models available for settling are well known and are generally based on mathematical concepts.
Most simulators use similar one-dimensional (ID) settling models for primary and secondary clarifiers.
The models may differ slightly in terms of the number of layers used to model the clarifier and the
methods for determining the settling velocities of particles. These models do not generally take into
account any biological reactions occurring in the clarifier or compression of the sludge blanket.
Recently, some models have integrated an activated sludge model to account for reactions taking place
in the clarifier.
One dimensional clarifier models typically are based on mass flux theory which calculates
settling in one dimension based on the sludge characteristics. One dimensional models cannot take into
account density currents and other non-vertical flows in a clarifier and thus represent an idealized
condition. In order to more properly represent actual clarifier performance 2 and 3 dimensional models
are necessary. These models have been constructed using computational fluid dynamics (CFD). CFD is a
mathematical modeling method which uses numerical techniques to solve the mass and energy
conservation equations in two and three dimensions. These models, while more complicated, can
model the currents caused by density differences and also flow currents due to baffling and other
structural features. A literature review of CFD techniques and studies can be found in Brouckaert and
Buckley (1998).
Activated Sludge
The most significant decision is in the selection of a model to describe the activated sludge
system. Activated sludge modeling first gained wide acceptance when the International Water
Association (IWA) released the activated sludge model, later known as ASM1. Although very
sophisticated, there are numerous simplifying assumptions underlying ASM1. These include constant
temperature and pH and no dependence of the biological reaction rates on food source. The model also
did not include phosphorus removal. ASM2 and ASM2d were developed to capture biological
phosphorus removal. ASM3 followed and includes a wider temperature range and allows for the effects
of different food sources. Additional equations can be added to ASM3 to account for biological
phosphorus removal. The metabolic biological phosphorus model of the Delft University of Technology
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(TUDP) was developed to fully account for the metabolism occurring in phosphorus accumulating
organisms (PAOs) during enhanced biological phosphorus removal. Barker and Dold (1997) developed a
model to include different rates of growth depending on the carbon source. The UCTPHO+ model is able
to model uptake of phosphorus under both anoxic and anaerobic conditions. Table C-l lists the
activated sludge models available for various suspended growth biological treatment processes.
Table C-l. Available Activated Sludge Models for Suspended Growth Processes
Model Name
ASMl
ASM2
ASM2d
ASMS
ASMS w/ BioP
TUDP
B&D
UCTPHO+
Wastewater Treatment Unit Processes
Carbon oxidation, nitrification, de-nitrification
Carbon oxidation, nitrification, de-nitrification, enhanced
biological phosphorus removal, fermentation, chemical
phosphorus removal
Carbon oxidation, nitrification, de-nitrification, enhanced
biological phosphorus removal, fermentation, chemical
phosphorus removal
Carbon oxidation, nitrification, de-nitrification
Carbon oxidation, nitrification, de-nitrification, enhanced
biological phosphorus removal
Carbon oxidation, nitrification, de-nitrification, enhanced
biological phosphorus removal, fermentation
Carbon oxidation, nitrification, de-nitrification, enhanced
biological phosphorus removal, fermentation
Carbon oxidation, nitrification, denitrification, biological
phosphorus removal
Reference
Henzeetal. 1987
Henzeetal. 1995
Henzeetal. 2000
Gujeretal. 1999
Reigeretal. 2001
Brdjanovic et al. 2000
Barker and Dold 1997
Huetal. 2007
Sources: WERF 2003, Gernaey et al. 2004, Comeau et al. 2008
Some reaction rates predicted by the ASM family of models for the liquid phase may be valid for
biofilm reactions as well, but the transfer of material to and from the biofilm for reaction cannot be
handled. Therefore, attached or fixed-growth processes require a different mathematical model to
represent transport of substrate to the biofilm, consumption of the substrate in the biofilm, and then
transport of the products out of the biofilm.
There are numerous mathematical models capable of modeling attached growth processes.
Unfortunately they are not as standardized as the activated sludge models and there are a considerable
number of them. As with activated sludge models, a model fully describing all aspects of the biofilm
process would be exceedingly complex and would require large amounts of computing power to solve.
Therefore numerous simplifying assumptions are made.
Biofilm models can be categorized as analytical, pseudo-analytical, 1-dimensional, and
multidimensional. In analytical models, enough assumptions are made that the equations can be solved
analytically. Psuedo-analytical models make a few less simplifying assumptions but require simple
numerical procedures to find a solution. One-dimensional models make less assumptions but must be
solved numerically, although they can still be solved using a desktop computer. Two and 3-dimensional
models are very complex and although they have fewer assumptions they require significantly more
computing time. These multi-dimensional models have not found wide use in industry and are not
included in any widely available simulators. They will not be discussed further. Table C.2 lists some
references available describing pseudo-analytical and 1-dimensional models.
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Table C-2. References for Biofilm Models
Model Type
Psuedo-Analytical
1-Dimensional
References
Rittmann and McCarty 2001, Grady et al. 1999, Saez and Rittmann 1992
Kissel et al. 1984, Wanner and Gujer 1986, Reichert 1998
In addition, work is continually being done to develop models for new and innovative processes.
For example, investigators have adapted biofilm models to simulate the Annamox process (Capuno et al.
2008; Dapena Mora et al. 2004). Models have also been developed to describe other emerging
processes such as SHARON (Volcke et al. 2006; Wett and Rauch 2003; Borger et al. 2008), DEMON and
STRASS (Wett et al. 2008). These models for new processes can be linked with the existing models for
activated sludge plants using simulators to model newer flow diagrams.
C.2 Comparison of Mathematical Models
Suspended Growth Activated Sludge Models
As summarized in Table C-l, there are a number of models available to represent activated
sludge processes. The list in C-l is not exhaustive, but only represents the major models in use today.
Each model has its own set of state variables. A state variable is a parameter used to describe the
condition of the system. In the case of activated sludge models they are the species that make up the
system such as particulate organic carbon and heterotrophic biomass. Each model is also represented
by a set of reactions which describe the transformations undergone by the state variables. Table C-3
summarizes the number of state variables and reactions each of the major models use.
Table C-3. State Variables and Reactions for Activated Sludge Models
Model
ASMl
ASM2
ASM2d
ASMS
ASM3w/BioP
TUDP
Barker and Dold
UCTPHO+
Number of State Variables
13
19
19
12
16
14
19
16
Number of Reactions
8
19
21
12
23
21
36
35
Source(s): Comeau and Takacs (2008), Gernaey et al, (2004)
While a complete list of all the state variables and reactions is beyond the scope of this
appendix, Table C-4 gives an example of the state variables used for the ASM3w/BioP model.
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Table C-4 State Variables for the ASM3w/BioP Model
State Variable
Ss
s,
So
Xs
XSTO
XPHA
X,
SNH
SNO
SNZ
Spo4
Xpp
XH
XA
XpAO
SHCO
XTSS
Description
Readily biodegradable organic matter
Soluble unbiodegradable organic matter
Dissolved oxygen
Slowly biodegradable organic matter
General storage compound
Stored PHA
Particulate unbiodegradable organic matter
Total ammonia
Total nitrate and nitrite
Nitrogen gas
Inorganic soluble phosphorus
Stored polyphosphate
Ordinary heterotrophic organisms
Nitrifying organisms
Phosphorus accumulating organisms
Soluble inorganic carbon
Total suspended solids
Units
mg COD/L
mg COD/L
mg 02/L
mg COD/L
mg COD/L
mg COD/L
mg COD/L
mgN/L
mgN/L
mgN/L
mgP/L
mgP/L
mg COD/L
mg COD/L
mg COD/L
mmol C/L
mg TSS/L
Source: Comeau and Takacs (2008)
The reactions of each model represent the transformation of the state variables. Comeau and
Takacs (2008) have developed a schematic representation of these reactions. Figure C-2 shows such a
schematic representation of the ASM3wBio/P model. Each line with an arrow represents a reaction in
the model. Each reaction is represented by constants which indicate the rate of the reaction and the
stoichiometry of the reaction. For example in the diagram the area proceeding from Ss to XSTO shows
that readily biodegradable matter in the presence of oxygen or nitrate is converted into general storage
compound in ordinary heterotrophic organisms. A reaction line meeting a vertical line outside the
dotted boundary represents the production of energy.
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.Ordinary heterotrophic
organisms
X
Nitrifying organisms
Phosphate Accumulating Organisms
JPO4
Figure C-2 Schematic Representation of ASMS w/BioP
Source: Comeau and Takacs 2008 from WEFTEC conference 2008. Reprinted with permission from Proceedings of WEFTEC®.08,
the 81st Annual Water Environment Federation Technical Exhibition and Conference, Chicago, IL, October 18-22, 2008.
Copyright © 2008 Water Environment Federation, Alexandria, Virginia, www.wef.org.
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Each model has its own set of reactions along with a set of assumptions underlying those
assumptions. Table C-5 shows the major models and some of the key assumptions. A discussion of each
follows the table.
Table C-5. Assumptions of Activated Sludge Models
Model
ASMl
ASM2
ASM2d
ASMS
ASMS w/ BioP
TUDP
Barker and Dold
UCTPHO+
Temperature
Dependent
N
Y
Y
Y
Y
Y
Y
Y
Microbial
Growth
DR
DR
DR
ER
ER
DR
DR
ER
Heterotroph
Growth Rate
Dependent on
Food Source
N
N
N
Y
Y
N
Y
Y
PAOs can
Denitrify
N
N
Y
N
Y
Y
Y
Y
PAO Decay
Dependent
on Food
Source
NA
N
N
NA
Y
Y
Y
Y
Key: DR - death-regeneration; ER- endogenous respiration; NA - not applicable.
Source: Gernaeyetal. 2004
Most reactions are temperature dependent. ASMl does not include temperature dependence,
instead including a choice of either 10 or 20 degrees Celsius. The other models do include temperature
dependent terms. Temperature dependent terms, however, are often limited in their applicability. For
example, the terms used in ASM2 are only valid between 10 and 25 degrees Celsius.
Microbial growth is generally described in one of two ways; death-regeneration or endogenous
respiration. The death-regeneration concept sets separate rates for the growth and decay of
microorganisms. Endogenous respiration assumes a maintenance concept where microorganisms use a
portion of their mass to provide energy. Each approach has its strengths and weaknesses. Death-
regeneration allows the model to take into account different reactions occurring upon death of the
microorganism and different fates of the material. It lumps together, however, the reactions of
heterotrophs and autotrophs. The endogenous respiration concept clearly separates heterotroph and
autotroph reactions but cannot take into account the different reactions occurring on the death of the
microorganisms.
Microbes can utilize some food sources more easily than others. For example, while both
methanol and VFA can be used as a substrate for denitrifying bacteria, only VFA can be used as a food
source for phosphorus accumulating organisms (PAO). Also bacterial growth rates can differ depending
on the carbon source. Many models do not take this into account.
It has been shown that denitrification can be accomplished by PAOs in the anaerobic zone of
biological phosphorus removal plants. Some of the biological phosphorus models include this option.
Important limitations of various models are listed below:
• ASMl models do not take into account either nitrite production or production of nitrogen gas.
Therefore it is not possible to close the nitrogen balance using the ASMl model.
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• The ASM2, ASM2d, and TUDP models do not account for GAO competition with PAOs during
biological phosphorus removal.
None of the models listed in Table C-2 can account for filamentous biomass or sludge bulking.
Biological phosphorus models cannot handle total depletion of substrate or the simultaneous presence
of substrate and another food source.
Biofilm Models
Characterization for the biofilm models are based on model type and assumptions about not
only kinetics and biological growth but also transport and the nature of the biofilm. Table C-6
summarizes each type of model and the assumptions and types of problems it can solve.
Table C-6. Applicability of Biofilm Models
Model Type
Analytical
Psuedo-Analytical
1-Dimensional
Assumptions
1 substrate is rate limiting, homogenous 1-D biofilm, first
or zero order kinetics
1 microbial species limited ability to model multiple
species, 1 rate limiting substrate, monod kinetics
Heterogenous biomass, mass transfer in biofilm
Analytical models are very simple and can only handle very simple problems. They are limited
to biofilms of even thickness and made up of a single species. Only a single rate limiting substrate can be
used. This does not mean that the reaction of more than one substrate cannot be handled, but that
microbial growth can only be limited by one of them. For example, consumption of both nitrate and
oxygen can be modeled, but the modeler must decide beforehand which will be the rate limiting
concentration for microbial growth. Analytical models can only use first or zero order rate constants
instead of the more realistic monod rate constants. Although no system matches these details exactly,
some may be close enough depending on the objective. If the only interest is in steady-state
consumption of a single substrate (e.g. nitrate) and the biofilm is fairly even, the analytical model may
be appropriate. Non-monod kinetics may be appropriate if the half saturation constant is not near the
concentration of substrate in the bulk solution. If the substrate concentration is much less than the half
saturation constant, first order kinetics can be used. If the substrate concentration is much greater than
the half saturation constant, zero order kinetics are appropriate. If the substrate concentration and half
saturation constant are about the same magnitude, analytical models are not appropriate.
Pseudo-analytical models use most of the same assumptions about biomass as the analytical
models. They still generally only take into account a single species and rate limiting substrate. Some
modifications can be made to handle multiple species if each species has only one limiting substrate.
The biggest advantage of the pseudo-analytical model is that it can include monod kinetics. Otherwise
its applicability is similar to the analytical model.
The 1-dimensional models keep separate track of dissolved and particulate components and can
model the concentration of a substrate throughout the biofilm with time. These models are able to use
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any kinetic expressions and can handle multiple species and multiple rate limiting substrates. They
cannot handle the effects of fluid velocity on substrate transfer or biofilm growth. They also cannot
handle uneven biofilm growth. The disadvantage to the 1-dimensional models is they require more
computing power and time, although many simulator programs today now include 1-dimensional
models.
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Brdjanovic, D., M.C.M. van Loosdrecht, P. Versteeg, CM. Hooijmans, G.J. Alaerts, and J.J. Heijnen.
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Brouckaert, C.J., and C.A. Buckley. 1998. Research on the Application of Computational Fluid Dynamics
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Autotrophic Nitrogen Removal under Different Reactor Configurations. Presented at WEFTEC 2008.
Comeau, Y. and I. Takacs. 2008. Schematic Representation of Activated Sludge Models. Presented at
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