United States Office of Research and EPA/625/R-95/008

Environmental Protection Development April 1996

Agency Cincinnati, OH 45268

1EPA Technology Transfer

Handbook

Management of Water

Treatment Plant Residuals
EPA/625/R-95/008

April 1996

Technology Transfer Handbook

Management of Water Treatment Plant Residuals

U.S. Environmental Protection Agency

Office of Research and Development

National Risk Management Research Laboratory

Cincinnati, Ohio

American Society of Civil Engineers

New York, New York

American Water Works Association

Denver, Colorado
Notice

The U.S. Environmental Protection Agency (EPA) through its Office of Research and Development

partially funded and collaborated in the research described here under EPA Contract 68-C3-0315

to Eastern Research Group, Inc. It has been subjected to the Agency’s peer and administrative

review and has been approved for publication as an EPA document.

The material presented in this publication has been prepared in accordance with generally recognized engineering principles and practices and is for general information only. This information

should not be used without first securing competent advice with respect to its suitability for any

general or specific application.

The contents of this publication are not intended to be a standard of the American Water Works

Association (AWWA) or the American Society of Civil Engineers (ASCE) and are not intended for

use as a reference in purchase specifications, contracts, regulations, statutes, or any other legal

document.

No reference made in this publication to any specific method, product, process, or service constitutes or implies an endorsement, recommendation, or warranty thereof by AWWA, ASCE, or EPA.

AWWA, ASCE, and EPA make no representation or warranty of any kind, whether expressed or

implied, concerning the accuracy, product, or process discussed in this publication and assume no

liability. Anyone using this information assumes all liability arising from such use, including but not

limited to infringement of any patent or patents.

Authorization to photocopy material for internal or personal use under circumstances not falling

within the fair use provisions of the Copyright Act is granted by ASCE to libraries and other users

registered with the Copyright Clearance Center (CCC) Transactional Reporting Service, provided

that the base fee of $4.00 per article plus $.50 per page is paid directly to CCC, 222 Rosewood

Drive, Danvers, MA 01923. The identification for ASCE Books is 0-7844-0181-0/96, $4.00 + $.50

per page. Requests for special permission or bulk copying should be addressed to Permissions &

Copyright Dept., ASCE.

Copyright © 1996 by the American Society of Civil Engineers and the American Water Works

Association. All rights reserved.

ii
Contents

Page

Chapter 1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1

1.1 Overview of Differences in Water Treatment Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . 1

1.2 Overview of Residuals Categories . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2

1.3 Overview of Residual Solids Treatment Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2

1.4 Selection of Residuals Management Plan Options . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4

1.5 Handbook User’s Guide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5

Chapter 2 Regulatory Issues Concerning Management of 

Water Treatment Plant Residuals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6

2.1 Discharge to Waters of the United States . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6

2.1.1 Technology-Based Effluent Limitations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6

2.1.2 Water Quality Standards . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7

2.1.3 Special Concerns Regarding Aluminum . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7

2.2 Discharge to Wastewater Treatment Plants. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7

2.3 Landfilling . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8

2.3.1 Municipal Nonhazardous Solid Wastes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8

2.3.2 Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9

2.4 Land Application . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9

2.4.1 Federal Regulations. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9

2.4.2 State Regulations. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10

2.5 Underground Injection. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10

2.5.1 Underground Injection Control Program . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10

2.5.2 Underground Injection Requirements Under RCRA . . . . . . . . . . . . . . . . . . . . . . . . 11

2.6 Disposal of Radioactive Waste. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12

2.7 Hazardous Waste . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12

2.8 Air Emissions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13

2.8.1 Federal Regulations. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13

2.8.2 State Regulations. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14

iii
Chapter 3 Characterization of Water Treatment Plant Residuals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17

3.1 Types and Quantities of Residuals. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17

3.1.1 Sludges . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17

3.1.2 Liquid Wastes. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18

3.1.3 Radioactive Wastes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 25

3.2 Physical Characteristics of Residuals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 30

3.2.1 Solids Content . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 31

3.2.2 Specific Resistance . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 33

3.2.3 Compressibility . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 34

3.2.4 Shear Stress . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 35

3.2.5 Density . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 36

3.2.6 Particle Size Distribution . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37

3.3 Chemical Characteristics of Residuals. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 39

3.3.1 Solids Content . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 39

3.3.2 Metals Content. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 39

3.3.3 Toxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 40

Chapter 4 Water Treatment Residuals Processing . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 41

4.1 Residuals Handling Process Types . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 41

4.2 Process Descriptions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 42

4.2.1 Collection Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 42

4.2.2 Thickening . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 43

4.2.3 Conditioning . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 45

4.2.4 Dewatering . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 46

4.2.5 Drying . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 52

4.2.6 Additional Residuals Handling Processes. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 53

4.3 Residuals Handling Process Performance. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 53

4.3.1 Process Performance. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 53

4.3.2 Comparison of Thickening Processes. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 54

4.3.3 Comparison of Dewatering Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 54

4.4 Developing Preliminary Residuals Processing Alternatives. . . . . . . . . . . . . . . . . . . . . . . . . 54

4.4.1 Residuals Processing Requirements. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 54

4.4.2 Preliminary Selection of Residuals Processing Alternatives. . . . . . . . . . . . . . . . . . 56

4.5 Specific Residuals Unit Process Selection Criteria . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 60

4.5.1 Operating Experience . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 60

4.5.2 Bench-Scale and Pilot-Scale Tests . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 61

Contents (continued)

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4.5.3 Environmental Impacts. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 62

4.6 Final Screening of Residuals Handling Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 62

4.6.1 Additional Selection Criteria. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 62

4.6.2 Process Alternative Selection Matrix . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 63

4.7 Residuals Handling Process Design Issues. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 63

4.7.1 Mass Balance Diagrams . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 63

4.7.2 Equipment Sizing . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 63

4.7.3 Contingency Planning . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 64

4.7.4 Specific Design Elements of Mechanical Dewatering Systems . . . . . . . . . . . . . . . 65

4.8 Air Emissions Control . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 68

4.8.1 Gaseous Residual Byproducts . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 68

4.8.2 Accidental Release of a Gaseous Treatment Chemical . . . . . . . . . . . . . . . . . . . . . 71

Chapter 5 Direct Discharge of Water Treatment Plant Residuals to Surface Waters . . . . . . . . . . . . . . 73

5.1 Theory . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 73

5.1.1 Chemical Interactions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 73

5.1.2 Toxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 74

5.2 Applications . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 77

5.2.1 Stream Hydraulics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 77

5.2.2 Available Transport and Chemical Models and Their Application . . . . . . . . . . . . . 79

5.3 Examples. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 80

5.3.1 California Plant, Cincinnati Water Works, Cincinnati, Ohio . . . . . . . . . . . . . . . . . . 80

5.3.2 Ralph D. Bollman Water Treatment Plant, Contra Costa Water 

District, California. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 83

5.3.3 Mobile Water Treatment Plant, Mobile, Alabama . . . . . . . . . . . . . . . . . . . . . . . . . . 86

5.3.4 City of Phoenix Utility, Phoenix, Arizona . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 88

5.4 Recommended Practices . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 100

Chapter 6 Discharge to Wastewater Treatment Plants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 102

6.1 Background . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 102

6.2 Survey of Operating Systems. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 102

6.3 Design Considerations and Conveyance Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 102

6.3.1 Regulatory Considerations. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 104

6.3.2 Conveyance System Design Considerations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 105

6.4 Design Considerations for Wastewater Treatment Plants . . . . . . . . . . . . . . . . . . . . . . . . . . 109

6.4.1 Hydraulic Loading . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 110

6.4.2 Organic Loading. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 110

Contents (continued)

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6.4.3 Solids Loading . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 110

6.4.4 Toxics Loading . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 111

6.4.5 Liquid/Solids Separation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 111

6.4.6 In-Plant Solids Handling . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 113

6.5 Ultimate Disposal of Wastewater Treatment Plant Biosolids. . . . . . . . . . . . . . . . . . . . . . . . 116

6.5.1 Direct Discharge. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 117

6.5.2 Land Application. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 117

6.5.3 Incineration. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 117

6.5.4 Composting . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 118

Chapter 7 Landfill Options . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 119

7.1 Landfill Siting. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 119

7.1.1 Airports . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 120

7.1.2 Floodplains . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 120

7.1.3 Wetlands . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 120

7.1.4 Fault Areas. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 120

7.1.5 Seismic Impact Zones . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 121

7.1.6 Unstable Areas. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 121

7.2 Landfill Design. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 121

7.2.1 Performance-Based Design Under 40 CFR Part 258. . . . . . . . . . . . . . . . . . . . . . . 121

7.2.2 Minimum Technology-Based Design Under 40 CFR Part 258 . . . . . . . . . . . . . . . . 122

7.3 Landfill Operations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 125

7.4 Metal Content Considerations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 125

7.4.1 Classification as Hazardous or Nonhazardous Waste . . . . . . . . . . . . . . . . . . . . . . 126

7.4.2 Mobility of Trace Metals. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 127

7.5 Dewatering . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 129

7.5.1 Mechanical Dewatering . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 130

7.5.2 Nonmechanical Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 130

7.6 Physical Characteristics of Water Sludges. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 130

7.6.1 Plasticity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 130

7.6.2 Compaction Behavior. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 131

7.6.3 Compressibility . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 131

7.6.4 Shear Strength. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 132

Chapter 8 Land Application . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 136

8.1 Regulatory Requirements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 136

8.2 Environmental Considerations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 136

Contents (continued)

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8.2.1 Major Components of Water Treatment Residuals and Their Impact on 

Soil Parameters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 137

8.2.2 Effects of Trace Metals Concentrations on Soil Properties . . . . . . . . . . . . . . . . . . 139

8.2.3 Impact of Water Treatment Residuals on the Availability of 

Phosphorus in Agricultural Soils . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 139

8.2.4 Effects of WTP Residuals on Soil Physical Properties. . . . . . . . . . . . . . . . . . . . . . 140

8.3 Land Application Options . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 141

8.3.1 Agricultural Land . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 141

8.3.2 Silviculture . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 143

8.3.3 Land Reclamation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 143

8.3.4 Dedicated Land Disposal. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 143

8.3.5 Other Use Options for WTP Residuals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 144

8.4 Operational Considerations in Land Application . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 144

8.4.1 Application Procedures . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 144

8.4.2 Public Participation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 144

8.4.3 Transportation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 144

8.4.4 Monitoring. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 145

Chapter 9 Brine Waste Disposal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 146

9.1 Background Information . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 146

9.1.1 Amount of Concentrate Generated and Disposal Methods . . . . . . . . . . . . . . . . . . 146

9.1.2 Constraints and Concerns . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 146

9.1.3 Early Disposal Regulations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 147

9.1.4 Current Regulations and Their Trends . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 147

9.2 Conventional Disposal Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 148

9.2.1 Surface Water Discharge. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 148

9.2.2 Disposal to Sanitary Sewers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 149

9.2.3 Deep Well Injection . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 150

9.2.4 Boreholes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 166

9.2.5 Spray Irrigation. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 167

9.3 Nonconventional Methods of Concentrate Disposal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 168

9.3.1 Evaporation and Crystallization of Brines for Zero Discharge . . . . . . . . . . . . . . . . 168

9.3.2 Evaporation Ponds. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 175

9.3.3 Emerging Technologies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 176

9.4 Costs Associated With Brine Waste Disposal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 176

9.5 Conclusion. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 176

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Chapter 10 Radioactive Waste Disposal. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 177

10.1 Background . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 177

10.2 Waste Disposal Practices . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 178

10.3 Waste Disposal Guidelines. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 178

10.3.1 Liquid Disposal. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 178

10.3.2 Solids and Sludge Disposal. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 180

10.4 Recordkeeping . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 181

Chapter 11 Economics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 182

11.1 Cost Assumptions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 182

11.1.1 Capital Cost Assumptions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 182

11.1.2 Operation and Maintenance Assumptions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 183

11.1.3 Total Annual Cost Assumptions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 183

11.1.4 Cost Components Excluded . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 183

11.1.5 Cost Equations. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 184

11.1.6 Cost Curves . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 184

11.1.7 Calculating Residuals Management Costs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 184

11.2 Gravity Thickening . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 185

11.2.1 Design Assumptions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 185

11.2.2 Capital Components. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 185

11.2.3 Operation and Maintenance Components. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 185

11.2.4 Cost Components Excluded . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 185

11.2.5 Gravity Thickening Cost Equations and Cost Curves. . . . . . . . . . . . . . . . . . . . . . . 185

11.3 Chemical Thickening. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 186

11.3.1 Design Assumptions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 186

11.3.2 Capital Components. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 186

11.3.3 Operation and Maintenance Components. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 187

11.3.4 Cost Components Excluded . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 187

11.3.5 Chemical Precipitation Cost Equations and Cost Curves. . . . . . . . . . . . . . . . . . . . 187

11.4 Mechanical Sludge Dewatering . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 188

11.4.1 Pressure Filter Press Cost Components. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 188

11.4.2 Scroll Centrifuge Cost Components . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 190

11.5 Nonmechanical Sludge Dewatering . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 191

11.5.1 Storage Lagoons Cost Components . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 191

11.5.2 Evaporation Ponds Cost Components . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 194

11.6 Discharge to Publicly Owned Treatment Works. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 196

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11.6.1 Design Assumptions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 196

11.6.2 Capital Components. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 196

11.6.3 Operation and Maintenance Components. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 196

11.6.4 Cost Components Excluded . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 196

11.6.5 Cost Equations and Cost Curves for Discharge to Publicly Owned 

Treatment Works . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 197

11.7 Direct Discharge . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 200

11.7.1 Design Assumptions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 200

11.7.2 Capital Components. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 201

11.7.3 Operation and Maintenance Components. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 201

11.7.4 Cost Components Excluded . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 201

11.7.5 Cost Equations for Direct Discharge. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 201

11.8 Land Application . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 203

11.8.1 Liquid Sludge Land Application Cost Components. . . . . . . . . . . . . . . . . . . . . . . . . 203

11.8.2 Dewatered Residuals Land Application Cost Components . . . . . . . . . . . . . . . . . . 204

11.9 Nonhazardous Waste Landfill. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 206

11.9.1 Off-Site Nonhazardous Waste Landfill Cost Components . . . . . . . . . . . . . . . . . . . 206

11.9.2 Onsite Nonhazardous Waste Landfill Cost Components . . . . . . . . . . . . . . . . . . . . 207

11.10 Hazardous Waste Landfill. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 209

11.10.1 Design Assumptions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 209

11.10.2 Cost Components . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 209

11.10.3 Cost Components Excluded . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 209

11.10.4 Total Annual Cost Equation and Cost Curves. . . . . . . . . . . . . . . . . . . . . . . . . . . . . 209

11.11 Radioactive Waste Disposal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 210

11.11.1 Low-Level Radioactive Waste Disposal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 210

11.11.2 Cost Information for Radioactive Waste Disposal. . . . . . . . . . . . . . . . . . . . . . . . . . 210

Chapter 12 Case Studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 211

12.1 Case Study 1: Disposal of Water Treatment Residuals From Pine Valley Water 

Treatment Plant, Colorado Springs, Colorado . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 211

12.1.1 Facility Information. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 211

12.1.2 Residuals Management . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 211

12.1.3 Residuals Handling Facilities and Operations. . . . . . . . . . . . . . . . . . . . . . . . . . . . . 212

12.1.4 Residuals Disposal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 215

12.1.5 Handling and Disposal Costs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 215

12.2 Case Study 2: Disposal of Water Treatment Residuals From Mesa Treatment Plant, 

Colorado Springs, Colorado . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 219

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12.2.1 Facility Information. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 219

12.2.2 Residuals Management . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 220

12.2.3 Residuals Handling Facilities and Operations. . . . . . . . . . . . . . . . . . . . . . . . . . . . . 220

12.2.4 Residuals Disposal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 223

12.3 Case Study 3: Land Application of Water Treatment Plant Residuals at 

Cobb County-Marietta Water Authority, Marietta, Georgia. . . . . . . . . . . . . . . . . . . . . . . . . . 224

12.3.1 Facility Information. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 225

12.3.2 Residuals Handling Facilities. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 230

12.3.3 Ultimate Disposal—Land Application. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 232

12.3.4 Disposal Costs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 234

12.4 Case Study 4: Treatment of Residuals at the Lake Gaillard Water Treatment Plant, 

North Branford, Connecticut . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 235

12.4.1 Facility Information. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 235

12.4.2 Lessons Learned . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 236

12.5 Case Study 5: Disposal of Water Treatment Plant Residuals From the San Benito 

Water Plant, Brownsville, Texas . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 236

12.5.1 Residuals Handling Facilities. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 237

12.5.2 Final Disposal. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 237

12.5.3 Disposal Costs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 237

12.6 Case Study 6: Management of Water Treatment Plant Residuals in the Chicago Area, 

Chicago, Illinois. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 237

12.6.1 Description of the Jardine Water Purification Plant . . . . . . . . . . . . . . . . . . . . . . . . 237

12.6.2 Residuals Handling and Disposal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 238

12.6.3 Disposal Costs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 239

Chapter 13 Waste Minimization and Reuse . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 242

13.1 Waste Minimization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 242

13.1.1 Process Modifications . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 242

13.1.2 Dewatering . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 243

13.1.3 Drying . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 243

13.2 Chemical Recovery. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 244

13.2.1 Coagulant Recovery. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 244

13.2.2 Lime Recovery . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 244

13.3 Innovative Use and Disposal Options . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 244

13.3.1 Beneficial Use . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 244

13.3.2 Disposal Options . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 245

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Chapter 14 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 246

Appendix A Survey of State Regulatory Requirements: Summary of Results. . . . . . . . . . . . . . . . . . . . . 254

Appendix B 1992 Survey of Water Treatment Plants Discharging to Wastewater Treatment Plants. . . 261

Appendix C Charges From Publicly Owned Treatment Works . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 268

Appendix D Chemical Monthly Average Doses, Pine Valley Water Treatment Plant, 

Colorado Springs, CO, 1987–1992 (Pine Valley, 1994) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 270

Appendix E Chemical Monthly Average Doses, Mesa Water Treatment Plant, 

Colorado Springs, CO, 1987–1992 (Mesa, 1994) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 276

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Figures

Figure Page

1-1 The primary target of a residuals management plan . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4

1-2 Handbook user’s guide. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5

2-1 Areas exceeding the ozone NAAQs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15

3-1 Generation of wastewater volumes with ion exchange . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21

3-2 Change in sludge settled solids concentration throughout a treatment plant . . . . . . . . . . . . . . . . . . . . . . . 32

3-3 Effect of Ca-to-Mg ratio on the solids concentration of softening sludge . . . . . . . . . . . . . . . . . . . . . . . . . . 32

3-4 Water distribution and removal in a softening (CaCO3) slurry and coagulant (Al(OH)3) slurry . . . . . . . . . . 33

3-5 Use of specific resistance to determine optimum chemical dosage . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 34

3-6 Variation in shear strength with sludge moisture content . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 35

3-7 Comparison of sludge settled solids concentration with the solids concentration where a sludge 

becomes “handleable” . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 36

3-8 Compaction curves of test sludges . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 36

3-9 Variation of floc density with floc size . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37

3-10 Variations in dewatered cake solids concentration of aluminum hydroxide sludges as a function of 

organic content . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37

3-11 Effect of incorporation of organic carbon on the relative size distribution of aluminum hydroxide 

sludge floc formed at pH 6.5 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37

3-12 Variations in measured floc density as a function of both coagulation pH and presence or absence of 

TOC from sludge floc matrix. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37

3-13 Floc size and resistance of metal hydroxide sludges to dewatering by vacuum filtration . . . . . . . . . . . . . . 38

3-14 Effect of specific surface area on the specific resistance of alum sludges . . . . . . . . . . . . . . . . . . . . . . . . . 38

3-15 Representative results from metal hydroxide sludge conditioning studies. . . . . . . . . . . . . . . . . . . . . . . . . . 39

3-16 Effect of Gt on optimum polymer dose for alum sludge conditioning . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 39

4-1 Residuals sources in water treatment plants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 42

4-2 Residuals handling process categories . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 42

4-3A Residuals handling process schematic: sedimentation basin used water flow . . . . . . . . . . . . . . . . . . . . . . 43

4-3B Residuals handling process schematic: solids dewatering . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 43

4-4 Gravity thickener cross-section. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 44

4-5 Gravity belt thickener cross section . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 46

4-6 Sand drying bed section . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 47

4-7 Wedgewire drying bed cross section . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 48

4-8 Dewatering lagoon cross section . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 48

4-9 Belt filter press . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 50

4-10 Solid-bowl-type centrifuge schematic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 50

4-11 Solid-bowl-type centrifuge . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 52

4-12 Vacuum filter . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 53

xii
4-13 Residuals handling process selection flow chart . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 60

4-14 Bench/pilot testing decision flow. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 61

4-15A Mass balance schematic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 64

4-15B Mass balance calculation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 65

4-16 Preliminary residuals handling process schematic. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 66

4-17 Aerial view of residuals handling process system, Val Vista WTP, Cities of Phoenix/Mesa, AZ . . . . . . . . . 67

4-18 Belt filter press—example of system layout . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 68

4-19 Centrifuge—example of system layout. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 68

4-20 Filter press—example of system layout . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 69

4-21 Vacuum filter—example of system layout . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 70

4-22 Packed tower . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 71

5-1 Flow frequency analysis of the Schuylkill River, Philadelphia, PA; minimum 7-day average flow 

values, 1932-1964 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 78

5-2 Scenarios for mass balance calculations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 78

5-3 Decay of a nonconservative pollutant. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 79

5-4 Control techniques for improving downstream water quality . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 80

5-5 Flow schematic of California Plant, Cincinnati Water Works, Cincinnati, OH . . . . . . . . . . . . . . . . . . . . . . . 81

5-6 Sediment aluminum concentration (dry weight) from Ohio River, Cincinnati, OH . . . . . . . . . . . . . . . . . . . . 82

5-7 Flow schematic of Ralph D. Bollman WTP, Contra Costa Water District, CA . . . . . . . . . . . . . . . . . . . . . . . 84

5-8 Sediment aluminum concentration from Mallard Reservoir, Concord, CA . . . . . . . . . . . . . . . . . . . . . . . . . 85

5-9 Flow schematic of Mobile WTP, Mobile, AL . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 86

5-10 Sediment aluminum concentration from Three Mile Creek, Mobile, AL . . . . . . . . . . . . . . . . . . . . . . . . . . . 87

5-11 Location of City of Phoenix WTPs, Phoenix, AZ . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 88

5-12 Recommended practices for direct discharge of WTP residuals to surface waters. . . . . . . . . . . . . . . . . . 100

6-1 Effect of WTP sludge on the combined volume of wastewater sludge after 30 minutes of settling . . . . . 112

7-1 Considerations for sludge monofill design . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 123

7-2A Compaction curve, Ferric Sludge 3 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 128

7-2B Compaction curve, Alum Sludge 1 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 128

7-2C Compaction curve, Alum Sludge 2 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 128

7-3 Consolidation curves of Alum Sludges 1 and 2 and Ferric Sludge 3 . . . . . . . . . . . . . . . . . . . . . . . . . . . . 131

7-4 Void ratio versus consolidation pressure of treated and untreated Alum Sludge 1 . . . . . . . . . . . . . . . . . . 132

7-5 Strength versus solids content for Alum Sludges 1 and 2 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 132

7-6 Strength versus curing time at various solids contents for ferric sludge . . . . . . . . . . . . . . . . . . . . . . . . . . 133

7-7 Shear strength versus additive level, Alum Sludge 2 (nonaged) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 134

7-8 Undrained sheer strength versus curing time for untreated and treated Alum Sludge 1 . . . . . . . . . . . . . . 134

7-9 Landfill height versus shear strength/unit weight for different slope angles . . . . . . . . . . . . . . . . . . . . . . . 135

8-1 Simplified planning procedure for land application of WTP residuals . . . . . . . . . . . . . . . . . . . . . . . . . . . . 137

8-2 Partitioning of trace metals in WTP residuals. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 139

8-3 Average (of three cuttings) phosphorus concentration in sorghum-sudangrass grown in Colby soil . . . . 140

9-1 Monthly operational report, page 1. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 158

9-2 Monthly operational report, page 2. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 159

Figures (continued)

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9-3 Monthly operational report, page 3. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 160

9-4 Monthly operational report, page 4. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 161

9-5 Monthly operational report, page 5. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 162

9-6 Well drilling cost . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 167

9-7 Zero discharge . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 168

9-8 Brine concentrator system . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 169

9-9 Brine concentrator capital and operating costs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 171

9-10 Brine concentrator cost components . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 173

9-11 Waste crystallizer system . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 173

9-12 Steam driven circulation crystallizer . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 174

9-13 Steam power crystallizer capital and operating costs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 175

9-14 Steam power crystallizer cost components . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 175

9-15 MVR crystallizer capital and operating costs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 175

9-16 MVR crystallizer cost components . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 176

11-1 Capital costs for gravity thickening. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 186

11-2 O&M costs for gravity thickening . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 186

11-3 Capital costs for chemical precipitation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 187

11-4 O&M costs for chemical precipitation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 187

11-5 Capital costs for pressure filter press . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 189

11-6 O&M costs for pressure filter press . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 189

11-7 Capital costs for scroll centrifuge . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 191

11-8 O&M costs for scroll centrifuge. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 191

11-9 Capital costs for lime softening storage lagoon . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 193

11-10 O&M costs for lime softening storage lagoon. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 193

11-11 Capital costs for alum sludge storage lagoon . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 194

11-12 O&M costs for alum sludge storage lagoon . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 194

11-13 Capital costs for evaporation ponds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 195

11-14 O&M costs for evaporation ponds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 195

11-15 Capital costs for 500 feet of discharge pipe . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 197

11-16 O&M costs for 500 feet of discharge pipe . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 197

11-17 Capital costs for 1,000 feet of discharge pipe . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 198

11-18 O&M costs for 1,000 feet of discharge pipe . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 198

11-19 Capital costs for 500 feet of discharge pipe with storage lagoon. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 199

11-20 O&M costs for 500 feet of discharge pipe with storage lagoon . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 199

11-21 Capital costs for 1,000 feet of discharge pipe with storage lagoon . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 200

11-22 O&M costs for 1,000 feet of discharge pipe with storage lagoon. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 200

11-23 Capital costs for liquid sludge land application. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 204

11-24 O&M costs for liquid sludge land application . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 204

11-25 Capital costs for trucking system . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 205

11-26 O&M costs for trucking system. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 205

11-27 Capital costs for dewatered sludge land application . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 206

Figures (continued)

Figure Page

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11-28 O&M costs for dewatered sludge land application . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 206

11-29 O&M costs for off-site nonhazardous waste landfill . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 207

11-30 Capital costs for onsite nonhazardous waste landfill . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 208

11-31 O&M costs for onsite nonhazardous waste landfill. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 208

11-32 Closure costs for onsite nonhazardous waste landfill . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 208

11-33 Postclosure costs for onsite nonhazardous waste landfill . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 208

11-34 O&M costs for hazardous waste disposal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 210

11-35 O&M costs for stabilization and hazardous waste disposal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 210

12-1 Locus map of Pine Valley WTP, Colorado Springs, CO . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 212

12-2 Schematic of Pine Valley WTP, Colorado Springs, CO . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 213

12-3 Flow chart of Pine Valley WTP, Colorado Springs, CO . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 214

12-4 Sediment and decant piping schematic, Pine Valley WTP, Colorado Springs, CO . . . . . . . . . . . . . . . . . . 215

12-5 Letter regarding new sludge handling facility at Pine Valley WTP, Colorado Springs, CO . . . . . . . . . . . . 216

12-6 Mesa WTP, Colorado Springs, CO . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 221

12-7 Letter from Colorado Springs Department of Utilities, CO . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 228

12-8 Schematic process diagram of Quarles WTP, Marietta, GA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 230

12-9 Schematic process diagram of Wyckoff WTP, Marietta, GA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 230

12-10 Schematic flow diagram of Jardine Water Purification Plant, Chicago, IL . . . . . . . . . . . . . . . . . . . . . . . . 238

12-11 Schematic diagram of residuals management system, Jardine Water Purification Plant, Chicago, IL . . . 239

Figures (continued)

Figure Page

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Tables

Table Page

1-1 Treatment Processes and Waste Streams . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3

1-2 Representative Solids Concentration Treatment Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4

2-1 Required Management Practices for Nonhazardous Industrial Waste Only and Co-Disposal Landfills, 

and Land Application Sites . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8

2-2 TCLP Constituents and Regulatory Limits . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13

2-3 National Ambient Air Quality Standards . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14

3-1 Alum/Iron Coagulant Sludge Characteristics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17

3-2 Chemical Softening Sludge Characteristics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17

3-3 Typical Chemical Constituents of Ion Exchange Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20

3-4 Regeneration of Cation Exchange Resins . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21

3-5 Membrane Process Operations Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22

3-6 Membrane Process Applications for RO, NF, and ED-EDR . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23

3-7 Concentration Factors for Different Membrane System Recoveries . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23

3-8 Tabulation of Concentration Factors. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 25

3-9 Summary of Treatment Processes and the Types of Wastes Produced From the Removal of 

Radionuclides From Drinking Water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26

3-10 Water Treatment Process Materials Containing Radionuclides . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26

3-11 Summary of Radium Concentration in Lime Softening Sludges and Backwash Water . . . . . . . . . . . . . . . 27

3-12 Summary of Radium-226 Concentrations in Brine Waste From Ion Exchange Treatment . . . . . . . . . . . . . 28

3-13 Summary of Uranium Concentrations in Ion Exchange Treatment Plant Wastewater. . . . . . . . . . . . . . . . . 28

3-14 Summary of Radium-226 Concentrations in Waste Stream From Iron Removal Filters . . . . . . . . . . . . . . . 28

3-15 Summary of Radium-226 Concentrations in Reject Water of Reverse Osmosis Treatment . . . . . . . . . . . . 29

3-16 Summary of Uranium Concentration in Reject Water of Reverse Osmosis Treatment . . . . . . . . . . . . . . . 29

3-17 Concentration of Radionuclides on Water Treatment Process Media and Materials . . . . . . . . . . . . . . . . . 30

3-18 Settled Solids Concentration of Residuals From Water Treatment Plants in Missouri . . . . . . . . . . . . . . . . 31

3-19 Effect of Coagulation Mechanism on Alum Sludge Properties . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 31

3-20 Specific Gravity of Sludge Particles and Cake Solids Concentrations Obtainable From Various 

Laboratory Dewatering Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 33

3-21 Specific Resistance for Various Chemical Sludges . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 34

3-22 Summary of Floc Density and Dewatered Solids Concentration Data for Several Chemical Sludges . . . . 35

4-1 Typical Ranges of Conditioner Use for Hydroxide Sludges in Various Mechanical 

Dewatering Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 46

4-2 Comparison of Thickening Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 54

4-3 Comparison of Dewatering Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 54

4-4 Preliminary Residuals Processing Selection Matrix . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 57

4-5 Survey of Thickening Methods at Water Treatment Plants in the United States . . . . . . . . . . . . . . . . . . . . 60

xvi
4-6 Survey of Dewatering Methods at Water Treatment Plants in the United States . . . . . . . . . . . . . . . . . . . . 60

4-7 Example of Weighting System for Alternative Analysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 64

4-8 Residual Handling Facility Contingency Planning Issues . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 67

5-1 Possible In-Stream Water Quality Guidelines and Standards . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 76

5-2 Average Daily Chemical Use at California Plant, Cincinnati Water Works, Cincinnati, OH . . . . . . . . . . . . 81

5-3 Chemical Composition of Sludge and Ohio River Water Sampled at California Plant, Cincinnati Water 

Works, Cincinnati, OH, September 21, 1988, December 18, 1988, and January 10, 1989 . . . . . . . . . . . . 82

5-4 S. Capricornutum Test Results on Alum Sludge Extracts From California Plant, Cincinnati Water 

Works, Cincinnati, OH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 82

5-5 Benthic Macroinvertebrates Collected From Site CO on the Ohio River, Cincinnati, OH, 1989 . . . . . . . . . 83

5-6 S. Capricornutum Test Results on Alum Sludge Extracts From Ralph D. Bollman Water Treatment 

Plant, Contra Costa, CA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 84

5-7 Benthic Macroinvertebrates Collected From Mallard Reservoir at Site CC, Contra Costa Water 

District, Concord, CA, February 21, 1989 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 85

5-8 S. Capricornutum Test Results on Alum Sludge Extracts From the Mobile, AL, Water 

Treatment Plant . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 87

5-9 Benthic Macroinvertebrates Collected From Three Mile Creek, Mobile, AL, 1989 . . . . . . . . . . . . . . . . . . . 87

5-10 Typical Canal Source Water Characteristics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 89

5-11 Summary of Plant Flows and Turbidity Data . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 90

5-12 Val Vista WTP Discharge Stream Characteristics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 90

5-13 Squaw Peak WTP Discharge Stream Characteristics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 90

5-14 Deer Valley WTP Discharge Stream Characteristics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 90

5-15 Existing Discharge Quantities . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 91

5-16 Grain Size Distribution for WTP Residuals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 91

5-17 Estimated Amount of Solids Deposited in the Canal System, Phoenix, AZ . . . . . . . . . . . . . . . . . . . . . . . . 91

5-18 Total Number of Fish Caught by Electrofishing, Arizona Canal, Phoenix, AZ, October 15, 1992, to 

April 30, 1993 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 92

5-19 Estimated WTP Discharge Stream Pollutant Concentrations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 93

5-20 Daily Resultant Pollutant Concentration Contribution to Canal at Typical Flow, Phoenix, AZ . . . . . . . . . . 94

5-21 Daily Resultant Pollutant Concentration Contribution to Canal at 25% of Typical Flow, Phoenix, AZ. . . . . 94

5-22 Toxicity of Arsenic to Freshwater Organisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 95

5-23 Toxicity of Cadmium to Freshwater Organisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 96

5-24 Toxicity of Chromium to Freshwater Organisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 97

5-25 Toxicity of Selenium to Freshwater Organisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 98

5-26 Toxicity of Mercury to Freshwater Organisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 99

5-27 Maximum Contaminant Levels (μg/L) Cold Water Fishery . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 99

6-1 Survey of Water Treatment Plants Discharging to WWTPs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 103

6-2 Treatment Chemical Analysis Range of Detected Contaminant Levels. . . . . . . . . . . . . . . . . . . . . . . . . . . 108

6-3 Comparison of Digester 11 With Background Digesters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 114

6-4 Alum Treatment Plant Sludge Dewatering Test Results . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 116

7-1 Solid Waste Landfill Criteria: Monofill for WTP Residuals and Co-disposal of Residuals With 

Nonhousehold Solid Waste. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 119

Tables (continued)

Table Page

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7-2 Solid Waste Landfill Criteria: Co-disposal of WTP Residuals With Municipal Solid Wastes . . . . . . . . . . . 119

7-3 Field Investigations for New Information, Landfill Design. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 122

7-4 Maximum Contaminant Levels in Uppermost Aquifer at Relevant Point of Compliance . . . . . . . . . . . . . 124

7-5 Comparison of Metals Concentrations in WTP Residuals, Natural Soils, and Sewage Sludge . . . . . . . . 126

7-6 Worst Case TCLP Results Using Maximum Metals Concentrations in Table 7-5, Compared With 

TCLP Regulatory Limits . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 127

7-7 EP Toxicity Test Results for Alum Residuals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 127

7-8 TCLP Results for Five Water Treatment Plant Coagulant Residuals. . . . . . . . . . . . . . . . . . . . . . . . . . . . . 127

7-9 Chemical Constituents of Synthetic Rainwater. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 128

7-10 Total Metals Analysis for Sludges Used in Leaching Research . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 129

7-11 Leaching of Metals in Lysimeter Test . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 129

7-12 Maximum Metals Concentrations in Lysimeter Leachate Compared With MCLs . . . . . . . . . . . . . . . . . . . 129

7-13 Liquid Limit, Plastic Limit, and Plasticity Index of Water Sludges . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 131

7-14 Shear Strength Parameters of Test Sludges . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 133

7-15 Remolded and Cured Undrained Strengths and Strength Gain Ratio . . . . . . . . . . . . . . . . . . . . . . . . . . . . 133

7-16 Undrained Shear Strength of Alum Sludge, Untreated and Treated . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 134

7-17 Required Shear Strength and Solids Concentration for Hypothetical Monofill Supporting Various 

Types of Heavy Equipment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 135

8-1 Composition of WTP Residuals Compared With Sewage Sludge and Agronomic Soils . . . . . . . . . . . . . . 138

8-2 Agronomic Components in WTP Residuals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 138

8-3 Phosphorus Recommendations for Several Agronomic Crops . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 143

9-1 Membrane Concentrate Generation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 146

9-2 Concerns and Requirements Associated With Conventional Disposal Methods . . . . . . . . . . . . . . . . . . . . 147

9-3 Concentrate Disposal Costs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 166

9-4 Typical Brine Concentrator Process Conditions in Zero Discharge Applications . . . . . . . . . . . . . . . . . . . . 169

9-5 Effect of Concentration Factor (CF) on Calcium Sulfate Seed Concentrations . . . . . . . . . . . . . . . . . . . . . 169

9-6 Typical Brine Crystallizer Process Conditions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 173

10-1 Summary of Residuals Produced From Water Treatment Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . 178

10-2 Water Treatment Methods for Residuals Containing Radionuclides . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 178

11-1 Capital Cost Factors and Selected Unit Costs for WTP Facility Planning . . . . . . . . . . . . . . . . . . . . . . . . 182

11-2 Operation and Maintenance Cost Factors and Unit Costs for WTP Facility Planning . . . . . . . . . . . . . . . 183

11-3 Flow Rate Use in Calculating Facility Costs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 184

11-4 Holding Tank Capacities, Gravity Thickening . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 185

11-5 Capital Cost Equation Determinants, Chemical Precipitation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 187

11-6 Capital Cost Equation Determinants, Pressure Filter Presses . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 189

11-7 Capital Cost Equation Determinants, Scroll Centrifuge . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 190

11-8 Capital Cost Equation Determinants, Lime Softening Storage Lagoons . . . . . . . . . . . . . . . . . . . . . . . . . 192

11-9 Capital Cost Equation Determinants, Alum Storage Lagoons . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 192

11-10 Capital Cost Equation Determinants, Evaporation Ponds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 195

11-11 Capital Cost Equation Determinants, Discharge to POTW . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 196

11-12 Capital Cost Equation Determinants, Direct Discharge . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 201

Tables (continued)

Table Page

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11-13 Capital Cost Equation Determinants, Liquid Residuals Land Application . . . . . . . . . . . . . . . . . . . . . . . . . 203

11-14 Capital Cost Equation Determinants, Dewatered Residuals Land Application . . . . . . . . . . . . . . . . . . . . . 205

12-1 Water Treated and Used, Pine Valley WTP, Colorado Springs, CO . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 215

12-2 Sediment Disposal, Pine Valley WTP, Colorado Springs, CO . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 219

12-3 Characteristics of Residuals Generated in 1982, Pine Valley WTP, Colorado Springs, CO . . . . . . . . . . . 220

12-4 Characteristics of Residuals Generated in 1992, Pine Valley WTP, Colorado Springs, CO . . . . . . . . . . . 221

12-5 Laboratory Samples: April 1992, Fountain Creek, Colorado Springs, CO . . . . . . . . . . . . . . . . . . . . . . . . 222

12-6 Laboratory Samples: August 1992, Mesa WTP, Colorado Springs, CO . . . . . . . . . . . . . . . . . . . . . . . . . 223

12-7 Raw Water Source for Mesa WTP, Colorado Springs, CO . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 224

12-8 Water Treated and Used, Mesa WTP, Colorado Springs, CO, in Million Gallons . . . . . . . . . . . . . . . . . . . 224

12-9A Chemical Characteristics of Residuals (Sludge) From Mesa WTP, Colorado Springs, CO, 1978. . . . . . . 225

12-9B Chemical Characteristics of Residuals (Supernatant) From Mesa WTP, Colorado Springs, CO, 1978 . . . 226

12-10 Chemical Characteristics of Residuals Generated in 1992, Mesa WTP, Colorado Springs, CO . . . . . . . 227

12-11 Sediment Disposal, Mesa WTP, Colorado Springs, CO . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 227

12-12 TCLP Data, Cobb County-Marietta Water Authority, Marietta, GA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 231

12-13 Total Metals Data, Cobb County-Marietta Water Authority, Marietta, GA . . . . . . . . . . . . . . . . . . . . . . . . . 231

12-14 Other Analyses, Cobb County-Marietta Water Authority, Marietta, GA . . . . . . . . . . . . . . . . . . . . . . . . . . . 232

12-15 Pesticides and PCBs (Solids), Cobb County-Marietta Water Authority, Marietta, GA . . . . . . . . . . . . . . . . 232

12-16 Triazine Herbicides, Cobb County-Marietta Water Authority, Marietta, GA . . . . . . . . . . . . . . . . . . . . . . . . 233

12-17 Reactivity, Solids, Cobb County-Marietta Water Authority, Marietta, GA . . . . . . . . . . . . . . . . . . . . . . . . . . 233

12-18 Pilot Study, Soil Data, Cobb County-Marietta Water Authority, Marietta, GA . . . . . . . . . . . . . . . . . . . . . . 234

12-19 Pilot Study, Plant Tissue Data, Cobb County-Marietta Water Authority, Marietta, GA . . . . . . . . . . . . . . . 234

12-20 Water Treatment Sludges Discharged by the City of Chicago WTPs to the District, Chicago, IL, 

1984–1992 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 239

12-21 Water Treatment Sludges Discharged by Various Suburban WTPs to the District, Chicago, IL, 

1984–1992 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 240

12-22 Local Limits on Dischargers Into District Sewarage Systems, Chicago, IL . . . . . . . . . . . . . . . . . . . . . . . . 240

12-23 Concentrations of Metals Found in WTP Sludges Discharged Into the District, Chicago, IL . . . . . . . . . . 240

12-24 User Charge Costs for Disposal of Water Treatment Residuals to the Metropolitan Water 

Reclamation District, Chicago, IL, 1984–1992 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 241

12-25 Water Treatment Sludges Discharged From Various Water Treatment Plants to the District From 

1984 Through 1992. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 241

C-1 Sewer Rates—Large Cities . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 268

C-2 Sewer Rates—Minnesota Cities. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 269

D-1 Raw Water, Pine Valley, 1987. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 270

D-2 Raw Water, Pine Valley, 1988. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 271

D-3 Raw Water, Pine Valley, 1989. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 272

D-4 Raw Water, Pine Valley, 1990. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 273

D-5 Raw Water, Pine Valley, 1991. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 274

D-6 Raw Water, Pine Valley, 1992. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 275

E-1 Raw Water, Mesa, 1992 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 276

Tables (continued)

Table Page

xix
E-2 Raw Water, Mesa, 1991 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 277

E-3 Raw Water, Mesa, 1990 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 278

E-4 Raw Water, Mesa, 1989 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 279

E-5 Raw Water, Mesa, 1988 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 280

E-6 Raw Water, Mesa, 1987 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 281

Tables (continued)

Table Page

xx
Conversion Factors

To convert . . . to . . . multiply by . . .

acres hectares 0.4046944

cubic feet cubic meters 0.02831685

degrees Fahrenheit degrees Celsius t°C = (t°F - 32)/1.8

feet meters 0.3048

inches centimeters 2.54

miles kilometers 1.609344

ounces grams 28.3495

pounds kilograms 0.45354237

pounds per 1,000 gallons grams per liter 0.1198322

pounds per square inch kiloPascals 6.895

square inches square centimeters 6.4516

tons metric tons 0.90718474

U.S. gallons liters 3.785

xxi
Acknowledgments

The organizations responsible for the development of this handbook are the U.S. Environmental

Protection Agency’s (EPA’s) National Risk Management Research Laboratory (NRMRL), Technology Transfer and Support Division, the American Society of Civil Engineers (ASCE), and the

American Water Works Association (AWWA). EPA’s Office of Research and Development partially

funded this effort under Contracts 68-C0-0068 and 68-C3-0315 with Eastern Research Group, Inc.

(ERG). Funding was also provided by ASCE.

Many individuals participated in the preparation and review of this handbook. Below is a partial list

of the major contributors.

Chapter 1: Introduction

Primary author: Jerry Russell, John Carollo Engineers, Phoenix, AZ. Contributing authors: Carl P.

Houck, Camp Dresser & McKee Inc., Denver, CO; Janine B. Witko, Malcolm Pirnie, Inc., Mahwah,

NJ; Brian Peck, John Carollo Engineers, Phoenix, AZ; and James E. Smith, Jr., EPA NRMRL,

Cincinnati, OH.

Chapter 2: Regulatory Issues Concerning Management of Water Treatment 

Plant Residuals

Primary author: Scott Carr, Black & Veatch, Charlotte, NC. Contributing authors: James E. Smith,

Jr., EPA NRMRL, Cincinnati, OH; Suzanna I. McMillan, Black & Veatch, Charlotte, NC; and Nancy

E. McTigue, Environmental Engineering & Technology, Inc., Newport News, VA. Key reviewers: Alan

Rubin and Robert Southworth, EPA Office of Water, Washington, DC.

Chapter 3: Characterization of Water Treatment Plant Residuals

Primary author: Timothy A. Wolfe, Montgomery Watson, Inc., Cleveland, OH. Contributing authors:

Mike Mickley, Mickley and Associates, Boulder, CO; John Novak, Virginia Polytechnic Institute and

State University, Blacksburg, VA; Leland Harms, Black & Veatch, Kansas City, MO; and Thomas J.

Sorg, EPA NRMRL, Cincinnati, OH.

Chapter 4: Water Treatment Residuals Processing

Primary author: Jerry Russell, John Carollo Engineers, Phoenix, AZ. Contributing authors: Brian

Peck, John Carollo Engineers, Phoenix, AZ; Tim Stephens, HDR Engineering, Inc., Phoenix, AZ;

Duncan Browne, WRC Engineers and Scientists, Huntingdon Valley, PA; Jeannette Semon, City of

Stamford WPCF, Stamford, CT; Betsy Shepherd, John Carollo Engineers, Phoenix, AZ; and Dale

D. Newkirk, EBMUD-Water Treatment, Oakland, CA.

Chapter 5: Direct Discharge of Water Treatment Plant Residuals to Surface Waters

Primary author: Dennis George, Tennessee Technological University, Cookeville, TN. Contributing

author: Christina Behr-Andres, University of Alaska, Fairbanks, AK.

xxii
Chapter 6: Discharge to Wastewater Treatment Plants

Primary author: Carl P. Houck, Camp Dresser & McKee Inc., Denver, CO. Contributing authors:

Janine B. Witko, Malcolm Pirnie, Inc., Mahwah, NJ; Timothy A. Wolfe, Montgomery Watson, Inc.,

Cleveland, OH; Richard Tsang, Camp Dresser & McKee Inc., Raleigh, NC; and Paula Murphy, ERG,

Lexington, MA.

Chapter 7: Landfill Options

Primary author: Nancy E. McTigue, Environmental Engineering & Technology, Inc., Newport News,

VA. Contributing authors: Daniel Murray, EPA NRMRL, Cincinnati, OH; and Mark Wang, The

Pennsylvania State University, University Park, PA.

Chapter 8: Land Application

Primary author: Bob Brobst, EPA Region 8, Denver, CO. Contributing authors: David R. Zenz and

Thomas C. Granato, Metropolitan Water Reclamation District of Greater Chicago, Chicago, IL. Key

reviewers: Robert Southworth and Alan Rubin, EPA Office of Water, Washington, DC; and Herschel

A. Elliott, The Pennsylvania State University, University Park, PA.

Chapter 9: Brine Waste Disposal

Primary author: William J. Conlon, Rust Environment and Infrastructure, Sheboygan, WI. Contributing authors: Thomas D. Wolfe, The Palmyra Group, Rough & Ready, CA; and William Pitt, Camp

Dresser & McKee Inc., Miami, FL.

Chapter 10: Radioactive Waste Disposal

Primary author: Thomas J. Sorg, EPA NRMRL, Cincinnati, OH. Key reviewers: Marc Parrotta, EPA

Office of Water, Washington, DC; Nancy E. McTigue, Environmental Engineering & Technology, Inc.,

Newport News, VA; and Darren Lytle, EPA NRMRL, Cincinnati, OH.

Chapter 11: Economics

Primary author: Bruce Burris, HDR Engineering, Inc., Irvine, CA. Contributing author: Christopher

Lough, DPRA, St. Paul, MN. Key reviewer: Janine B. Witko, Malcolm Pirnie, Inc., Mahwah, NJ.

Chapter 12: Case Studies

Primary author: Balu P. Bhayani, Colorado Springs Utilities, Colorado Springs, CO. Contributing

authors: James M. Parsons, Cobb County-Marietta Water, Marietta, GA; Cecil Lue-Hing, Metropolitan Water Reclamation District of Greater Chicago, Chicago, IL; and Jerry Russell, John Carollo

Engineers, Phoenix, AZ.

Chapter 13: Waste Minimization and Reuse

Primary author: Janine B. Witko, Malcolm Pirnie, Inc., Mahwah, NJ. Contributing author: James E.

Smith, Jr., EPA NRMRL, Cincinnati, OH.

Peer Reviewers

James H. Borchardt, Montgomery Watson, Inc., Walnut Creek, CA; C. Michael Elliott, Stearns &

Wheeler, Cazenovia, NY; Terry L. Gloriod, Continental Water Company, St. Louis, MO; Paul E.

Malmrose, Montgomery Watson, Inc., Saddle Brook, NJ; Howard M. Neukrug, Philadelphia Water

Department, Philadelphia, PA; S. James Ryckman, University of Dayton, Dayton, OH; James K.

Schaefer, Metcalf & Eddy, Inc., Somerville, NJ; J. Edward Singley, Montgomery Watson, Inc.,

xxiii
Gainesville, FL; Kevin L. Wattier, Metropolitan Water District, Los Angeles, CA; and Thomas L. Yohe,

Philadelphia Suburban Water, Bryn Mawr, PA.

Handbook Development Team

Project directors: James E. Smith, Jr., EPA NRMRL, Cincinnati, OH; and Marla Berman, ASCE,

Washington, DC. Additional coordinators: Jon DeBoer, AWWA, Denver, CO; Nancy E. McTigue,

Environmental Engineering & Technology, Inc., Newport News, VA; and Janine B. Witko, Malcolm

Pirnie Inc., Mahwah, NJ. Technical and copy editing and production support were provided by

Eastern Research Group, Inc., of Lexington, Massachusetts.

xxiv
Chapter 1

Introduction

Potable water treatment processes produce safe drinking water and generate a wide variety of waste products

known as residuals, including organic and inorganic

compounds in liquid, solid, and gaseous forms. In the

current regulatory climate, a complete management program for a water treatment facility should include the

development of a plan to remove and dispose of these

residuals in a manner that meets the crucial goals of

cost effectiveness and regulatory compliance. The development of a comprehensive water treatment residuals management plan typically involves each of the

following steps:

• Characterize the form, quantity, and quality of the

residuals.

• Determine the appropriate regulatory requirements.

• Identify feasible disposal options.

• Select appropriate residuals processing/treatment

technologies.

• Develop a residuals management strategy that meets

both the economic and noneconomic goals established for a water treatment facility.

This handbook provides general information and insight

into each of the above-mentioned steps that a potable

water treatment facility should follow in developing a

residuals management plan. For additional information

on some of the processes described in this handbook,

see the Handbook of Practice: Water Treatment Plant

Waste Management (Cornwell et al., 1987) or Slib,

Schlamm, Sludge (Cornwell and Koppers, 1990).

1.1 Overview of Differences in Water

Treatment Processes

The form taken by water treatment plant (WTP) residuals can vary greatly, depending on the source of untreated water and the type of unit processes

incorporated in a water treatment facility. Although many

different forms of residuals are generated in the potable

water treatment industry, this handbook primarily addresses those residuals produced by the following general categories of WTPs, which are identified by the

names of their most common unit processes:

• Coagulation/Filtration plant: This traditional form of

WTP is typically used to remove turbidity and pathogenic organisms. These facilities may also be used

to remove color, taste, and odor-causing compounds

from the water supply. A variation of the process may

use aeration and oxidation processes for the removal

of iron and manganese. Unit processes may include

screening, chemical pretreatment, presedimentation,

microstraining, aeration, oxidation, coagulation/flocculation, sedimentation/precipitation, filtration, disinfection, and dissolved air flotation treatment. Other

nonchemical variations include direct filtration, diatomaceous earth filtration, and slow sand filtration.

• Precipitative softening plant: This variation of a coagulation/filtration facility uses additional processes

to reduce water hardness. Unit processes may include screening, chemical pretreatment, presedimentation, microstraining, aeration, oxidation, coagulation/

flocculation, lime softening, sedimentation/precipitation, filtration, and disinfection.

• Membrane separation: This process is typically used

to remove turbidity, total dissolved solids, hardness,

nitrates, and radionuclides from a water supply. More

recent applications address removal of microbiological contaminants. Membrane separation generally

involves the use of microfiltration, ultrafiltration, nanofiltration, reverse osmosis, or electrodialysis, often times

in combination with pretreatment practices.

• Ion exchange (IX): These facilities are used to remove inorganic constituents, including hardness, nitrates, arsenic, and radionuclides from water. The

process involves the use of IX reactors in combination with pretreatment processes.

• Granular activated carbon (GAC) adsorption: GAC is

used in many processes for the removal of naturally

occurring and synthetic organic matter from water.

The fundamental differences between the unit processes of these five plant types characterize the type of

residuals generated at a given facility. WTPs constructed to meet the ever-widening scope of future regulations may require a combination of these unit

processes—a coagulation/filtration treatment plant using

1
GAC as a filter medium, for example. Consequently, this

handbook’s discussion of different water treatment

methods is purposely broad to include existing types of

residuals as well as those from future potable water

treatment facilities. The treatment of off-gas from an air

stripping process and the residuals from diatomaceous

earth filtration are some of the residual streams not

extensively addressed in this handbook.

1.2 Overview of Residuals Categories

WTP residuals are typically derived from suspended

solids in the source water, chemicals (e.g., coagulants)

added in the treatment processes, and associated process control chemicals (e.g., lime). Some potable water

treatment processes produce residuals that are relatively straightforward to process and dispose of. For

example, the leaves, limbs, logs, plastic bottles, and

other large floating debris separated from water during

the screening process are simply disposed of at conventional solid waste landfills. Most other treatment processes generate more complex residual waste streams

that require more sophisticated processing methods and

final disposal methods to protect human health and the

environment. The four major categories of residuals

produced from water treatment processes are:

• Sludges (i.e., water that contains suspended solids

from the source water and the reaction products of

chemicals added in the treatment process). Presedimentation, coagulation, filter backwashing operations, lime softening, iron and manganese removal,

and slow sand and diatomaceous earth filtration all

generate sludge.

• Concentrate (brines) from IX regeneration and salt

water conversion, membrane reject water and spent

backwash, and activated alumina waste regenerant.

• IX resins, spent GAC, and spent filter media (including sand, coal, or diatomaceous earth from filtration

plants).

• Air emissions (off-gases from air stripping, odor control units, ozone destruction on units).

The chemical characteristics and contaminant concentration levels in these residual waste streams often dictate the ultimate disposal options. Furthermore, it is

reasonable to expect that as drinking water quality is

increasingly regulated, higher removal efficiencies of

more contaminants will be required. To achieve these

higher efficiencies, WTPs will need to use more advanced treatment technologies. Of potential concern is

the case where the residuals are characterized as either

hazardous or radioactive waste. Depending on the raw

water quality and treatment process removal efficiency,

hazardous or radioactive characteristics could be exhibited in potentially any residual waste stream mentioned

above. Classification of WTP residuals as hazardous or

radioactive material, however, is unlikely at this time or

in the near future; no water treatment residuals have yet

been classified as either.

Some other typical residual waste streams associated

with a WTP are considered beyond the scope of this

handbook. These waste streams include storm drainage, sanitary sewage flows, laboratory and building floor

drainage, and waste from spill containment areas, all of

which generate residuals processing and disposal concerns for the drinking water industry. Each of these

residual waste streams has specific regulatory requirements typically associated with point and nonpoint

source pollution control as defined under the Clean

Water Act (CWA). As with all residual treatment and

disposal issues, consult with the appropriate state and

local regulators to determine specific requirements.

Table 1-1 lists some typical residuals generated from

drinking water treatment processes, possible contaminants normally found in each waste stream, available

residual disposal methods, and a quick reference to the

applicable federal regulations. In all cases, specific state

and local regulations must also be considered.

1.3 Overview of Residual Solids

Treatment Processes

In many instances, regulatory requirements or the need

for cost effectiveness dictate that a residual receive

further treatment to make it acceptable for disposal. The

three classic treatment processes for residuals solids

are thickening, dewatering, and drying. Application of a

particular process depends on the solids concentration

of the residual. This handbook describes applications of

several types of solids concentrating processes.

Table 1-2 indicates which common residual treatment

processes are usually applied to residuals with low,

medium, or high solids concentrations, respectively.

Within the water treatment industry, the definitions of

low, medium, and high solids concentrations vary, depending on whether a sludge is produced by a coagulation/filtration plant or a precipitative softening plant. The

processes presented in Table 1-2 are often configured

in series to create combination systems that can provide

a low level of operational complexity with a high degree

of operational flexibility.

This handbook gives background information about

each of the most common treatment processes. This

information can then be used in the preliminary selection

and sizing of appropriate unit processes during the development of a residuals management plan. This handbook also briefly discusses available treatment

technologies for gaseous residuals. These technologies

include stripping, odor control, gaseous chemical leak

treatment, and ozonation.

2
Table 1-1. Treatment Processes and Waste Streams (Robinson and Witko, 1991)

Major

Treatment

Process

Type

Typical Residual Waste

Streams Generated

Typical Contaminant

Categories Typical Disposal Methods

Relevant

Chapter in

Handbook

Regulation Covering

Disposal Method

Coagulation/

Filtration

Aluminum hydroxide,

ferric hydroxide, or

polyaluminum chloride,

sludge with raw water

suspended solids,

polymer and natural

organic matter

(sedimentation basin

residuals)

Metals, suspended

solids, organics,

biological, radionuclides,

inorganics

Landfilling Chapter 7 RCRA/CERCLA

Disposal to sanitary

sewer/WWTP

Chapter 6 State and local

regulations

Land application Chapter 8 RCRA, DOT

Surface discharge Chapter 5 NPDES (CWA), state

and local DOH

Spent backwash

filter-to-waste

Metals, organics,

suspended solids,

biological, radionuclides,

inorganics

Recycle State and local DOH

Surface discharge

(pumping, disinfection,

dechlorination)

Chapter 5 NPDES (CWA), state

and local regulations

Disposal to sanitary

sewer/WWTP

Chapter 6 State and local

regulations

Precipitative

softening

Calcium carbonate and

magnesium hydroxide

sludge with raw water

suspended solids and

natural organic matter

Metals, suspended

solids, organics,

unreacted lime,

radionuclides

Landfilling Chapter 7 RCRA/CERCLA, state

and local regulations

Disposal to sanitary

sewer/WWTP

Chapter 6 State and local

regulations

Land application Chapter 8 RCRA, state and

local regulations, DOT

Spent backwash

filter-to-waste

Metals, organics,

suspended solids,

biological, radionuclides,

inorganics

Recycle Chapter 5 State and local DOH

Surface discharge

(pumping, disinfection,

dechlorination)

Chapter 6 NPDES (CWA), state

and local regulations

Disposal to sanitary

sewer/WWTP

State and local

regulations

Membrane

separation

Reject streams

containing raw water

suspended solids

(microfiltration), raw

water natural organics

(nanofiltration), and

brine (hyperfiltration, RO)

Metals, radionuclides,

TDS, high molecular

weight contaminants,

nitrates

Surface discharge

(pumping, etc.)

Chapter 9 RCRA, NPDES, state

and local regulations

Deep well injection

(pumping)

Chapter 9 RCRA, NPDES, state

and local regulations

Discharge to sanitary

sewer/WWTP

Chapter 9 State and local

regulations

Radioactive storage Chapter 10 RCRA, DOT, DOE

Ion exchange Brine stream Metals, TDS, hardness,

nitrates

Surface discharge Chapter 9 RCRA, NPDES, state

and local regulations

Evaporation ponds Chapter 9 RCRA, NPDES, state

and local regulations

Discharge to sanitary

sewer/WWTP

Chapter 9 State and local

regulations

Granular

activated

carbona

Spent GAC requiring

disposal and/or

reactivation, spent

backwash, and

gas-phase emissions in

reactivation systems

VOCs, SOCs

(nonvolatile pesticides),

radionuclides, heavy

metals

Landfill Chapter 7 RCRA, CERCLA, DOT

Regeneration—on/off site Chapter 14 State and local air

quality regulations

(CAA)

Incineration State and local air

quality regulations

(CAA)

Radioactive storage

Return spent GAC to

supplier

Chapter 10 DOT, DOE

3
1.4 Selection of Residuals Management

Plan Options

Before selecting the treatment and disposal actions necessary to develop a residuals management plan, the

manager of a WTP may start with a large array of

residuals processing and disposal options. These options are narrowed through consideration of specific

residuals characteristics and associated regulatory requirements. A focus on the available disposal options

further narrows the array to a finite set of residuals

management alternatives.

Figure 1-1 illustrates that the primary target during the

development of a residuals management plan should be

practical disposal options and treatment processes that

will take into account economic and noneconomic factors of concern to the community. The technical criteria

used in the selection of the final management plan differ

from user to user; economic, cultural, social, and environmental factors are also site-specific, and are typically

included in any final selection. The technical information

Major

Treatment

Process

Type

Typical Residual Waste

Streams Generated

Typical Contaminant

Categories Typical Disposal Methods

Relevant

Chapter in

Handbook

Regulation Covering

Disposal Method

Stripping

process

(mechanical

or packed

tower)

Gas phase emissions VOCs, SOCs, radon Discharge to atmosphere

GAC adsorption of off-gas

(contaminant type and

concentration dependent)

Not addressed State and local air

quality regulations

(CAA)

Spent GAC if used for

gas-phase control

VOCs, SOCs,

radionuclides

GAC adsorption of off-gas

(contaminant type and

concentration dependent)

Return spent GAC to

supplier

Not addressed State and local air

quality regulations

(CAA)

a The discussion on disposal methods for GAC residuals is generic in nature. For more specific information on disposal options for GAC, see

McTigue and Cornwell (1994).

Key

CAA = Clean Air Act

CERCLA = Comprehensive Environmental Response, Compensation and Liability Act

DOE = Department of Energy

DOH = Department of Health

DOT = Department of Transportation

NPDES = National Pollutant Discharge Elimination System

RCRA = Resource Conservation and Recovery Act

RO = Reverse osmosis

SOC = Synthetic organic chemical

TDS = Total dissolved solids

VOC = Volatile organic compound

Table 1-1. Treatment Processes and Waste Streams (Robinson and Witko, 1991) (Continued)

Table 1-2. Representative Solids Concentration Treatment Processesa

Process

Solids

Concentration Equalization

Gravity

Settling

Dissolved Air

Flotation Lagoon Mechanical Open Air

Thermal

Drying

Thickening Low X X X X X

Dewatering Medium X X X

Drying High X X X X

a Chapter 4 of this handbook provides a more complete discussion of processing alternatives.

Type of Water


Treatment Plant

Residual Characteristics

Regulatory


Requirements

Treatment


Options

Disposal


Options

• Social/Cultural


• Environmental


• Economic

Figure 1-1. The primary target of a residuals management plan.

4
in this handbook can be used to screen out inappropriate residuals processing and disposal options. Cost

curves for various treatment processes and a matrix of

commonly encountered social, cultural, and environmental factors are included.

1.5 Handbook User’s Guide

This handbook offers background information about the

components of, and the development process for, a

comprehensive residuals management plan. Developing a successful residuals management plan requires an

understanding of the value of residuals characterization

and the regulatory requirements, tailoring the treatment

options to the requirements of the available disposal

alternatives, and then developing rational evaluation criteria. This handbook is organized to logically guide the

user through each of these steps.

Figure 1-2 illustrates a generic decision process for

developing a residuals management plan. Each step in

the process is keyed to a chapter in this handbook.

Figure 1-2 can also be used as a quick reference to find

specific information throughout the book.

Water Treatment Plant Residuals


Addressed in This Handbook

Determine Characteristics

See Chapter 3

See Chapter 9

See Chapter 2

See Chapter 7 See Chapter 8

See Chapter 13

Is Residual a Brine?

See Chapter 10 Is Residual Radioactive?

Is


Stream Discharge


Feasible?

Identify Disposal Options

Is


Sewer Discharge 


Feasible?

See Chapter 5

See Chapter 8

Landfill Land Application

Identify Regulatory Constraints

Emerging Technologies Develop Residuals Processing


Treatment Alternatives

See Chapter 4

Select Cost-Effective


Residuals Handling Strategy

See Chapter 11

See Chapter 13

CASE


STUDIES

Yes

Yes

Yes

Yes

No

No

No

No

Figure 1-2. Handbook user’s guide.

5
Chapter 2

Regulatory Issues Concerning Management of 

Water Treatment Plant Residuals

Water treatment utility managers who are beginning to

explore alternative methods for disposal of water plant

wastes may encounter difficulty identifying the regulations that affect the various management practices. The

difficulty is compounded by the many different types of

wastes produced by water treatment plants (WTPs).

This chapter provides an overview of the regulatory

requirements governing the following disposal methods

for WTP residuals: direct discharge, discharge to wastewater treatment plants, disposal in landfills, land application, underground injection, disposal of radioactive

waste, and treatment of air emissions. Applicable federal regulations and typical state requirements are both

discussed. In addition, the results of two surveys of state

regulatory requirements conducted by the American

Water Works Association (AWWA) Water Treatment Residuals Management Committee and the American Society of Civil Engineers (ASCE), are summarized in

Appendix A. The surveys also identify commonly used

residuals management practices in each state.

At the federal level, EPA has not established any regulations that are specifically directed at WTP residuals.

Applicable regulations are those associated with the

Clean Water Act (CWA); Criteria for Classification of

Solid Waste Disposal Facilities and Practices (40 CFR

Part 257); the Resource Conservation and Recovery Act

(RCRA); the Comprehensive Environmental Response,

Compensation, and Liability Act (CERCLA); and the

Clean Air Act (CAA). The CWA limits direct discharges

into a water course, while the other legislation governs

other methods of use and/or disposal of wastes. Most

states are responsible for establishing and administering regulations that will meet the requirements of these

acts. The regulation of wastes, therefore, is primarily the

responsibility of the states.

2.1 Discharge to Waters of the 

United States

The federal program to protect the quality of the nation’s

water bodies is authorized under the Federal Water

Pollution Control Act (FWPCA) of 1972. Since its passage, the statute has been amended in 1977, 1978,

1980, 1981, and 1987, and renamed the Clean Water

Act (CWA). The act and associated regulations attempt

to ensure that water bodies maintain the appropriate

quality for their intended uses, such as swimming, fishing, navigation, agriculture, and public water supplies.

The National Pollutant Discharge Elimination System

(NPDES) applies to WTPs that discharge wastes directly to a receiving water. Under Section 402 of the

CWA, any direct discharge to waters of the United

States must have an NPDES permit. The permit specifies the permissible concentration or level of contaminants in a facility’s effluent. EPA authorizes states to act

as the primary agent for the NPDES program, provided

that the state program meets all EPA requirements.

State regulations and guidelines controlling the discharge of residuals, however, vary throughout the

United States. Certain states permit direct discharge of

residuals with or without pretreatment requirements. In

other states, direct discharge has been restricted

through limitations on suspended solids and pH (W.E.

Gates and Associates, 1981; Cornwell and Koppers,

1990). Generally, direct discharge into streams has

been permitted for clarified water such as settled backwash water or overflow from solids separation processes.

For states not granted primacy, EPA regional offices

issue NPDES permits. An NPDES permit is issued to a

discharger based on technology-based effluent limitations, water quality standards, or both.

2.1.1 Technology-Based Effluent Limitations

Under Sections 301 and 304 of the CWA, EPA is required to establish national effluent limitations for the

major categories of industrial dischargers. These limitations reflect the capabilities of the best available technology that is economically feasible for use in treating

industrial discharges to surface waters. Federal effluent

limitations have not yet been issued for WTP residuals;

therefore, the delegated states or the regional EPA offices are responsible for establishing the limits for WTP

discharges. Federal guidelines for controlling WTP discharges were drafted but never fully implemented. The

6
draft guidance document divides WTPs into three categories (W.E. Gates and Associates, 1981; Cornwell and

Koppers, 1990):

• Plants that use coagulation, oxidation for iron and

manganese removal, or direct filtration.

• Plants that use chemical softening procedures.

• Plants that use a combination of the procedures in

the above categories.

The draft document defines the best practical control

technology for each category and establishes discharge

limits on pH and total suspended solids (TSS). Based

on plant capacity, WTP discharge limits varied in the

document from 0.6 to 1.3 g TSS per cubic meter of water

treated, which corresponds to a TSS of approximately

30 to 60 mg/L in the discharge, assuming a 2 percent

waste discharge of the main stream. Because secondary treatment or the equivalent is required for treating

drinking water, 85 percent removal and/or a TSS level

of 30 mg/L are typically required for WTP discharges.

2.1.2 Water Quality Standards

Under Section 303 of the CWA, each state is required

to establish ambient water quality standards for its water

bodies. These standards define the type of use and the

maximum permissible concentrations of pollutants for

specific types of water bodies. The states use water

quality criteria documents published by EPA, as well as

other advisory information, as guidance in setting maximum pollutant limits. EPA reviews and approves the

state standards.

According to EPA, certain in-stream water quality standards at the edge of the mixing zone must be met to allow

direct discharge of WTP residuals. A controlled release

of water clarifier residuals and filter backwash that

meets water quality standards may be considered technology-based controls in appropriate circumstances

(Cornwell and Lee, 1993).

Some states have established maximum allowable concentrations for pollutants in discharges to water bodies.

These limits generally apply if they are more stringent

than the allowable discharge that will meet the in-stream

water quality criteria. For example, Illinois does not allow

a discharge of greater than 15 mg/L fluoride, and barium

discharge must be lower than 2 mg/L, even if the 1 mg/L

in-stream standard could be met through dilution in the

mixing zone (Cornwell et al., 1987). An overflow from a

solids separation process such as a lagoon, thickener,

or, sometimes, backwash water, requires a discharge

permit.

2.1.3 Special Concerns Regarding Aluminum

EPA has developed an ambient water quality standard

for aluminum, requiring that the in-stream soluble aluminum level not exceed 87 μg/L on a 4-day average. The

1-hour average must not exceed 750 μg/L. This standard may be of concern for WTPs that use aluminum

sulfate or other aluminum coagulants and rely on an

NPDES-permitted discharge to dispose of settled backwash wastewater, supernatant from dewatering process

units, or coagulation/flocculation residuals. The standard is focused on dissolved aluminum; most aluminum

in WTP discharges is in solid form.

Because aluminum is abundant in some geologic formations, it is not unusual to find it in concentrations higher

than 1 mg/L in receiving streams for NPDES discharges.

Consequently, some states may mandate complete

elimination of any NPDES discharges that contribute to

the aluminum load of the receiving stream.

2.2 Discharge to Wastewater Treatment

Plants

Three general categories of regulations affect the decision of whether to discharge WTP residuals to a sanitary

sewer or directly to a wastewater treatment plant

(WWTP): 1) federal and state hazardous waste regulations; 2) federal and state radioactive waste regulations;

and 3) local receiving wastewater utility regulations,

driven generally by the need for a utility to comply with

the provisions of the CWA.

The first two categories of regulation usually are not a

problem. WTP residuals are rarely, if ever, classified as

hazardous waste. While water treatment facilities often

generate a radioactive component in residuals, it is at a

very low level. The radioactive component results from

removing normally occurring radioactive material

(NORM) from the process water by the addition of a

coagulant. The resulting low-level radioactive WTP residuals are generally acceptable for disposal to a wastewater utility. The third category of regulations, local

receiving wastewater utility discharge permit regulations, incorporates any provisions concerning hazardous or radioactive wastes that may be germane.

Provisions are sometimes included that reduce the risk

of operational problems at the receiving WWTP or of

violation to its discharge permit. In small utility situations, often no formal local receiving wastewater utility

regulations exist, and compliance with any regulatory

guidance is of little concern.

All WWTPs must comply with EPA’s general pretreatment requirements, under which WTP residuals are

clearly identified as industrial wastes. In addition, EPA

requires all WWTPs with wastewater inflow greater than

5 million gallons per day (mgd) to establish formal pretreatment enforcement programs. Under these programs, the WWTPs regulate discharges into municipal

sewers by industries, including water treatment facilities,

and other entities. The WWTPs must, at a minimum,

7
enforce national pretreatment standards. They also may

implement additional controls or limits based on local

conditions for any of the following reasons:

• A WWTP’s NPDES permit may require removal of

pollutants that the plant itself cannot remove.

• Pollutants in residuals discharged into the sewer system may adversely affect the performance of the collection system or removal systems in the WWTP.

• Pollutants in residuals discharged into the sewer system may contaminate WWTP wastes, making dewatering and disposal more difficult and expensive.

EPA authorizes the states to review and approve individual WWTP pretreatment programs. In addition to the

individual plant’s pretreatment requirements, state

agencies may provide specific guidelines.

2.3 Landfilling

Under RCRA Subtitle D regulations (40 CFR Parts 257

and 258), criteria have been established for the design

and operation of nonhazardous, solid waste landfills.

Landfills that receive only drinking water treatment residuals are subject to the requirements of 40 CFR Part

257. These requirements are also applicable to landfills

that accept solid waste other than household waste,

such as industrial waste. These criteria apply to all

nonhazardous, nonhousehold solid waste disposal facilities and practices. They address seven areas of environmental concern pertaining to landfill design and

operation (see Table 2-1). It should be noted that these

criteria are performance based and do not include any

specific design criteria.

2.3.1 Municipal Nonhazardous Solid Wastes

Municipal solid waste landfills (MSWLFs) are subject to

the criteria of 40 CFR Part 258. Unlike Part 257, Part

258 includes specific design criteria in addition to performance-based criteria. If a utility disposes of its WTP

residuals in a monofill, then, under federal regulations,

Part 258 does not apply; Part 257 (discussed above)

does instead. If, however, the WTP residuals are co-disposed of with municipal solid waste, the requirements

established for MSWLFs apply. These landfill criteria

(40 CFR Part 258) address six major areas as listed in

Table 7-2.

Six location restrictions apply to MSWLFs. Landfills cannot be located in floodplains, wetlands, seismic impact

zones, and unstable areas. Additionally, specific setbacks are required for landfills near airports and fault

areas.

The minimum design criteria for new landfills are intended to give owners/operators two basic design options—a composite liner design, or a site-specific design

Table 2-1. Required Management Practices for Nonhazardous

Industrial Waste Only and Co-Disposal Landfills,

and Land Application Sites (40 CFR Part 257)

Environmental

Concern Management Practice

Floodplains Facilities in floodplains shall not restrict the flow of

the 100-year flood, reduce the temporary water

storage capacity of the floodplain, or result in

washout of solid waste, so as to pose a hazard to

human life, wildlife, or land or water resources.

Endangered

species

Facilities shall not cause or contribute to the taking

of any endangered or threatened species of plants,

fish, or wildlife, and shall not result in the

destruction or adverse modification of the critical

habitat of endangered or threatened species.

Questions about the potential for adversely

affecting endangered or threatened species at a

particular site should be directed to the nearest

Regional office of the U.S. Fish and Wildlife

Service. (Note: Notices of draft NPDES permits

also are routinely sent to the U.S. Fish and Wildlife

Services. See 40 CFR 124.10(c).)

Surface water Facilities shall not cause a discharge of pollutants

into waters of the United States in violation of

Section 402 of the Clean Water Act, shall not

cause a discharge of dredged or fill material in

violation of Section 404 of the Act, and shall not

cause nonpoint source pollution that violates an

approved Section 208 water quality management

plan.

Ground water Facilities shall not contaminate an underground

drinking water source beyond the solid waster

boundary, or beyond an alternative boundary.

Consult the regulation for procedures necessary to

set alternative boundaries (see Appendix D). To

determine contamination, Appendix I of 40 CFR

Part 257 provides a list of contaminant

concentrations. Release of a contaminant to

ground water which would cause the

concentrations of that substance to exceed the

level listed in Appendix I constitutes contamination.

Disease Disease vectors shall be minimized through the

periodic application of cover material or other

techniques as appropriate to protect public health.

Air Facilities shall not engage in open burning of

residential, commercial, institutional, or industrial

sold waste. Infrequent burning of agricultural

wastes in the field, silvicultural wastes for forest

management purposes, land-clearing debris,

diseased trees, debris from emergency cleanup

operations and ordinance is allowed.

Safety 1. Explosive gases generated by the facility shall

not exceed 25 percent of the lower explosive

limit for the gases in facility structures and 100

percent of the lower explosive limit at the

property line.

2. Facilities shall not pose a hazard to the safety of

persons or property from fires.

3. Facilities within 10,000 feet of any airport

runway used by turbojet aircraft, or within 5,000

feet of any airport runway used by piston-type

aircraft shall not pose a bird hazard to aircraft.

4. Facilities shall not allow uncontrolled public

access so as to expose the public to potential

health and safety hazards at the disposal site.

8
that meets the performance standard of RCRA Subtitle

D and has been approved by the permit writer.

Subtitle D establishes specific criteria for the following

landfill operational factors: regulated hazardous waste

detection program, cover material, disease vector control, explosive gases control, air criteria, access requirements, storm water run-on and runoff management,

surface water, liquids restrictions, recordkeeping.

The most extensive portion of Subtitle D pertains to

ground-water monitoring and corrective action. The rule

mandates that all landfills must have a monitoring system that yields sufficient ground-water samples from

around the site. Both upgradient and downgradient wells

and sampling are required. Tests to detect 15 heavy

metals and 47 volatile organic compounds must be

conducted, but approved states can modify the list to

reflect site-specific conditions.

The financial assurance requirements of RCRA are designed to guarantee that funds will be available for

closure and postclosure care of landfills and, if needed,

corrective action. All landfills are required to install a final

cover system with at least two components, an infiltration layer and an erosion layer. The minimum final cover

requirements mandate an infiltration layer at least 18

inches thick with a very low permeability limit, and an

erosion layer at least 6 inches thick that can sustain

native plant growth. A written closure plan must be

submitted to the state solid waste director. Postclosure

care must be provided for 30 years; it includes maintaining the final cover, the leachate collection and disposal

system, the ground-water monitoring program, and the

landfill gas monitoring system.

WTP residuals that are co-disposed of with municipal

solid waste in a sanitary landfill do not have to meet

numerical pollutant limits. The residuals, however, cannot be hazardous and must not contain free liquids.

Mechanical and natural dewatering of residuals is often

adequate for removing the free liquid from the water

treatment solids. Because EPA requires landfill operators to make random inspections of incoming waste or

take other steps to ensure that incoming loads do not

contain regulated hazardous wastes, landfill operators

may require residuals disposers to prove their solids are

not hazardous. Hazardous wastes (as defined in 40

CFR Part 261) are waste materials that exhibit ignitability, corrosivity, reactivity, or toxicity (see Section 2.7).

EPA now prohibits the disposal of noncontainerized or

bulk liquid wastes in landfills. Subtitle D requires that the

landfill owner or operator determine if the wastes (including municipal water treatment solids) are liquid

wastes according to the Paint Filter Liquids Method

9095. This simple test is performed by placing a representative sample of the solids in a mesh Number 60

paint filter (available at paint stores) and allowing it to

drain for 5 minutes. The solids are considered a liquid

waste if any liquid passes through the filter during the

5-minute period.

2.3.2 Summary

In summary, neither 40 CFR Part 257 nor Part 258

requires that WTP plant residuals be stabilized or have

a certain percent solids concentration. 40 CFR Part 258

does require that the residuals pass a paint filter test

prior to co-disposal with solid wastes in a landfill. Both

regulations require that the filled areas be periodically

covered to protect public health. Both regulations require attention to protection of ground and surface water

sources and control of gas migration.

Some states may impose further restrictions on bulk or

special wastes such as those produced during water

treatment. Besides the paint filter test, states also require a minimum total solids content. Nebraska, one of

the most restrictive states, requires a minimum total

solids content of 70 percent. Some states also have a

maximum allowable ratio of residuals-to-municipal solid

waste.

2.4 Land Application

2.4.1 Federal Regulations

Land application of WTP residuals is typically regulated

at the state level. The recently developed federal standards for the use or disposal of sewage residuals (40

CFR Part 503) specifically exclude WTP residuals. Criteria for classification of solid waste disposal facilities

and practices (40 CFR Part 257), however, could affect

land application of WTP residuals (see Table 2-1). This

rule regulates the disposal of nonhazardous wastes,

which include residuals generated from a WTP. The

objective of this rule is to prevent construction or operation of a residuals processing or disposal facility from

adversely affecting surface water, ground water, endangered or threatened wildlife, or public health. Criteria

were established under 40 CFR Part 257 regulating

application of residuals containing cadmium and polychlorinated biphenyls (PCBs). As discussed in Chapter

3 on residuals characterization, concentrations of cadmium and PCBs in WTP residuals are usually below

detection levels, making these criteria inapplicable.

Specific requirements for controlling disease vectors are

also included under Part 257 to prevent adverse public

health impacts resulting from application of solid

wastes/residuals. Criteria are also established to protect

ground-water quality beyond the application site boundary. Land application activities must not cause specific

organic and inorganic chemicals to exceed maximum

contaminant levels in the area. These criteria generally

should not affect the ability to land-apply WTP residuals.

9
2.4.2 State Regulations

Most state agencies are waiting for EPA to establish land

application criteria for WTP residuals. Despite the differences in their characteristics, many states apply WWTP

biosolids criteria, such as metals loading limits, to the

land application of WTP residuals.

Some states have taken it on themselves to establish

specific criteria for land application of WTP residuals.

For example, the Colorado Department of Health regulates beneficial land application of WTP residuals under

domestic sewage sludge regulations (Colorado, 1986).

The department has established specific regulations

pertaining to the beneficial use of WTP residuals on

land. These regulations require development of an approved beneficial use plan that:

• Identifies where the material will be used.

• Contains approval from the land owner and the appropriate county health department.

• Contains residuals analyses.

• Identifies the types of crops to be grown and the

application rates.

Parameters to be analyzed include aluminum, arsenic,

cadmium, pH, total solids, and nutrients. If WTP residuals are used with biosolids, then the practice must comply with the biosolids regulations. Application to land

where root crops or low-growing crops are to be grown

is prohibited if the crops are intended for human consumption.

The Colorado Department of Health does not allow

beneficial use of WTP residuals with radioactivity levels

exceeding 40 picocuries total alpha activity per gram

(pCi/g) of dry residuals.

The Missouri Department of Natural Resources has also

established guidelines for land application of alum residuals. These guidelines recommend that soil pH be

maintained near 7.0 for alum residuals application and

that total aluminum loading to the soil not exceed 4,000

lb/acre without site-specific investigations. No restrictions apply to land application of lime softening residuals

(Missouri, 1985).

2.5 Underground Injection

The federal government has promulgated regulations

for injecting wastes into the ground. Most states have

fully adopted these regulations and have been granted

primacy for enforcing the program. These regulations

alone, however, will probably not prevent underground

injection of WTP residuals. Some states, such as Florida, have chosen to promulgate and/or issue stricter

regulations and/or policies, making it very difficult to

dispose of WTP residuals such as brines via underground injection.

2.5.1 Underground Injection Control Program

Underground injection may be a disposal option for

concentrates and brines from drinking water treatment

processes. This option is subject to regulatory approval

under the underground injection control (UIC) program,

authorized by the Safe Drinking Water Act (SDWA). The

UIC program regulates the subsurface placement of

fluid in wells or dug-holes with a depth greater than their

width. The program covers the disposal of hazardous

waste and various other substances in wells.

After the 1980 amendments to the SDWA, EPA developed a mechanism to grant individual states primary

enforcement responsibility for underground injection.

The UIC regulations are enforced through a permitting

system. Each state is required to develop a UIC permit

program that enforces EPA standards, prevents underground injection unless authorized by a permit or rule,

authorizes underground injection only where the process will not endanger drinking water sources, and maintains thorough records and inspection reports. Currently,

all 50 states have established UIC programs, some of

which are managed by EPA regional offices.

2.5.1.1 Classification of Underground Injection

Wells

To prevent contamination of drinking water sources, EPA

established regulatory controls and a classification system based on the type of waste injected and the location

of the injection well (40 CFR 146.5):

• Class I includes wells used to inject hazardous waste,

or industrial and municipal waste beneath the lowermost formation containing an underground source of

drinking water within 1/4 mile of the well bore.

• Class II includes wells used to inject fluids generated

from oil and natural gas production and refining.

• Class III includes wells that are injected with liquid

(such as water) for extraction of minerals including

sulfur, uranium, and other metals in situ.

• Class IV includes wells that are used to inject hazardous or radioactive wastes into or above a formation containing an underground source of drinking

water within 1/4 mile of the well. This class of wells

also includes wells used to dispose of hazardous

waste into or above a formation containing an aquifer

that has been exempted pursuant to 40 CFR 146.04.

• Class V includes the injection wells that are not covered in Classes I through IV.

Wells used for underground injection of WTP residuals

fall into Classes I, IV, and V.

10
2.5.1.2 Permit Application

Underground injection wells are authorized by rule under RCRA and permitted at the state level under the

SDWA.

Authorization by Rule

Authorization of underground injection by rule gives

owners and operators an opportunity to operate underground injection facilities before their permit applications

are approved. To be authorized by rule, owners or operators must comply with the applicable requirements.

For Class I and III wells, the owners or operators must

meet the requirements listed in 40 CFR 144.28 and

individual state requirements. The authorization is void

if a permit application has not been filed in a timely

manner as specified in 40 CFR 144.31(c)(1); otherwise,

it can be extended until the permit is issued or revoked

by a UIC program administrator. For Class IV wells, the

operation period can be as long as 6 months on the

condition that requirements specified in 40 CFR 144.13,

144.14(c) are met. Currently, the operation of Class V

wells is not authorized by rule.

Authorization by Permit

Except for owners or operators who are authorized by

rule to run underground injection facilities, all other facilities must be authorized by permits. Facilities authorized by rule are also required to apply for a permit to

continue operating their facilities on a long-term basis.

All applicants for underground injection permits must

complete the application forms provided by a state UIC

program, and submit necessary supporting documents.

This paperwork should include the following information:

• Facility name, address, and ownership.

• Activities that are conducted in the facility and which

require a permit under RCRA, CWA, or CAA.

• A list of principal products or services provided by the

facility.

• Lists of the relevant permits and construction approvals issued to the facility.

• Geographic and topographic characteristics of the facility.

2.5.1.3 Underground Injection Control Criteria

and Standards

The UIC program administrator reviews the permit applications based on UIC criteria and standards (40 CFR

146). Different criteria and standards apply to different

classes of injection wells. The major criteria and standards are:

• Construction requirements: Configuration of the injection wells (hole size, depth of injection zone), injection

pressure, type of cement, and type of injected fluids

must be specified and meet the requirements. Both

new and existing wells must be cased and cemented

to protect sources of drinking water.

• Operating requirements: Certain operating conditions

must be met. For example, the injection pressure

must not exceed a maximum, and injection between

the outermost well casing and the well bore must be

avoided to protect underground sources of drinking

water.

• Monitoring requirements: The nature of the injected

fluids, injection pressure, flow rate, cumulative volume, and mechanical integrity must be monitored.

Monitoring should take place at regular intervals and

is based on class of well and operation type.

• Reporting requirements: Quarterly or yearly reports

should include the operating conditions for the period,

the results of the monitoring, and the results of any

other required tests.

2.5.1.4 Actions Against Violations

Under the SDWA, no operator or owner of an underground injection facility is allowed to construct, operate,

maintain, convert, plug, or abandon that facility, or conduct any other injection activity in a way that might

contaminate a ground-water source of drinking water.

For Class I, II, and III wells, if ground-water quality

monitoring shows the intrusion of contaminants to a

ground-water source, then corrective action may be required. Corrective actions may be taken in the areas of

operation, monitoring, or reporting, or might include

closing the well, if required. If the operation of the well

is authorized by a permit, modification of the permit with

additional requirements might occur (40 CFR 144.39). If

the permit is violated, appropriate enforcement action

might be taken (40 CFR 144.55).

Under the SDWA, new construction of most Class IV

wells is strictly prohibited. Increasing the amount or

changing the type of waste injected in these wells is also

forbidden.

If a Class V well causes a violation of primary drinking

water regulations under 40 CFR Part 142, the operator

of the well must take any action (including closing the

injection well) to prevent contaminating drinking water

sources.

2.5.2 Underground Injection Requirements

Under RCRA

Section 3004(f) of RCRA requires EPA to determine

whether underground injection of hazardous wastes will

endanger human health and the environment. In response, EPA has banned the underground injection of

hazardous wastes that do not meet the applicable treat11
ment standards of the land disposal restrictions (see 40

CFR Part 148). The 1986 Hazardous and Solid Waste

Amendments (HSWA) to RCRA enhanced the restriction of underground injection of hazardous waste. These

amendments prohibit the disposal of hazardous waste

through underground injection into or above a formation

within 1/4 mile of an underground source of drinking

water.

2.6 Disposal of Radioactive Waste

When radium isotopes and other radioactive materials

are removed from a drinking water supply, they are

usually concentrated in residuals that must be disposed

of in ways that are cost-effective, practical, and protective of human health and the environment. The U.S.

Nuclear Regulatory Commission (NRC) has the authority to regulate the handling and disposal of all licensed

synthetic radioactive material. Certain materials, such

as water and WWTP wastes containing naturally occurring radioactive material (NORM), are not licensed or

regulated by NRC because they are not included under

its authority as source, special nuclear, or byproduct

material. Regulation of NORM is left to individual states

(Hahn, 1989). Limitations on radioactive residuals are

established case-by-case, using best professional judgment (Koorse, 1993a). Recognizing that water treatment

could concentrate radionuclides, EPA has issued guidelines for managing radionuclide waste from water treatment containing up to 2,000 pCi/g (dry weight) of NORM

(U.S. EPA, 1994b). The guidelines addressing disposal

of residuals, brines, and solid wastes containing radium

or uranium are discussed in detail in Chapter 10.

Several states regulate disposal of WTP residuals containing radionuclides. For example, in certain states,

water quality standards include specific criteria for radionuclides. Some states also limit the discharge of

wastes containing naturally occurring radionuclides into

sanitary sewers. Illinois limits the use of residuals for soil

conditioning on agricultural lands. Illinois and Wisconsin

have developed criteria for landfilling residuals containing radium. In Wisconsin, the discharge of WTP residuals to a sanitary sewer collection and treatment system

that is otherwise acceptable is governed by radium discharge limits as follows:

CRa−226

400 +

CRa−228

800 ≤ 1

where

CRa-226 = the concentration in picocuries per liter

(pCi/L) of soluble Ra-226 in the wastewater

CRa-228 = the concentration of soluble Ra-228 in

pCi/L in the wastewater

The average monthly combined radium concentrations

discharge, measured in pCi/L, must not exceed the

limits of the preceding equation, and the total amount of

radiation released to the sanitary sewer system in any

1-year period cannot exceed 1.0 Ci (Hahn, 1989).

In Colorado, if the total alpha activity of the WTP residuals exceeds 40 pCi/g (dry weight), the WTP is required

to contact the Colorado Department of Health, Radiation

Control Division, for further disposal guidance (Colorado, 1990). In addition, North Dakota has regulations

requiring the water plant to be licensed as a generator

of radioactive material (Cornwell et al., 1987). Other

states have adopted similar regulations that may affect

whether WTP residuals containing radioactivity may be

discharged to sanitary sewers.

2.7 Hazardous Waste

WTP residuals are generally not considered hazardous

wastes. Even spent granular activated carbon (GAC)

residuals are usually not hazardous wastes (McTigue

and Cornwell, 1994; Dixon, 1993). Some disposal or use

measures for WTP residuals, however, require demonstration that the material is not hazardous according to

governing hazardous waste regulations.

Subtitle C, Section 3001 of RCRA addresses treatment,

storage, and disposal of hazardous waste. These regulations are designed to ensure proper management of

hazardous waste from “cradle to grave”—i.e., from the

moment the waste is generated until it is ultimately

disposed of. This approach has three elements:

• A tracking system requiring that a uniform manifest

document accompany any transported hazardous

waste from the point of generation to the point of final

disposal.

• An identification and permitting system that enables

EPA and the states to ensure safe operation of all

facilities involved in treatment, storage, and disposal

of hazardous waste.

• A system of restrictions and controls on the placement of hazardous waste on or into the land.

EPA employs two separate mechanisms for identifying

hazardous wastes (as defined in 40 CFR Part 261).

Wastes may be defined as hazardous based on their

characteristics (ignitability, corrosivity, reactivity, or toxicity), or they may be specifically designated as hazardous in lists published by the agency. Since WTP

residuals are not specifically designated as hazardous

wastes, they can be classified as hazardous only if they

exhibit any of the four hazardous characteristics.

A waste is classified as ignitable if it is capable of

causing a fire that burns vigorously and persistently

under standard pressure and temperature, causing a

hazard. An aqueous waste stream is corrosive if it has

a pH lower than 2.0 or higher than 12.5. A waste is

reactive if it has any of these characteristics:

12
• Is unstable.

• Reacts violently or forms potentially explosive mixtures with water.

• Generates toxic gases or fumes when mixed with water.

• Contains cyanides or sulfides that can generate toxic

gases or fumes.

• Can detonate if exposed to heat.

• Is already defined as an explosive material.

Although domestic WTP residuals are rarely ignitable,

corrosive, or reactive, the greatest concern for water

treatment wastes is toxicity, as determined by the Toxicity Characteristic Leaching Procedure (TCLP) (U.S.

EPA, 1992c). The TCLP applies to certain metals, herbicides, pesticides, and volatile organic compounds

(VOCs). A water treatment residual failing the TCLP test

can be classified as a hazardous material. Many landfills

require the disposer to document TCLP analysis results

as proof that the material is not toxic before solids

disposal is allowed.

The TCLP uses a vacuum-sealed extraction vessel to

capture VOCs in the sample. Table 2-2 lists the contaminants analyzed in the TCLP, along with maximum allowable pollutant concentrations at which the waste is

considered nonhazardous. WTP solids generally do not

fail the TCLP and are not classified as hazardous materials. If WTP wastes are classified as hazardous wastes,

they are subject to the handling, treatment, and disposal

requirements specified in RCRA Subtitle C.

CERCLA also affects landfilling of WTP wastes. Through

CERCLA, cleanup costs at hazardous waste sites can

be assessed against the user of the site on a volumeuse basis; the waste itself need not have directly caused

the problems. Co-disposal of water treatment residuals,

then, with potentially hazardous wastes significantly increases the potential for a utility to be held responsible

for landfill remediation costs.

2.8 Air Emissions

The federal government has promulgated regulations to

control air emissions from industrial sources. Most

states have fully adopted these regulations and have

been granted primacy for enforcing them. Some states

such as New Jersey, Michigan, and California, have

promulgated stricter regulations and/or policies governing air emissions. These states may require treatment

of air stripper emissions using activated carbon treatment or combustion.

2.8.1 Federal Regulations

Federal regulations do not specifically regulate air emissions from drinking water treatment plants. Instead, they

Table 2-2. TCLP Constituents and Regulatory Limits (40 CFR

Part 261.24)

Constituents Reg. Level (mg/L)

Arsenic 5.0

Barium 100.0

Benzene 0.5

Cadmium 1.0

Carbon tetrachloride 0.5

Chlordane 0.03

Chlorobenzene 100.0

Chloroform 6.0

Chromium 5.0

o-Cresol 200.0

m-Cresol 200.0

p-Cresol 200.0

Cresola 200.0

2,4-D 10.0

1,4-Dichlorobenzene 7.5

1,2-Dichloroethylene 0.5

1,1-Dichloroethylene 0.7

2,4-Dinitrotoluene 0.13

Endrin 0.02

Heptachlor (and its hydroxide) 0.008

Hexachlorobenzene 0.13

Hexachloro-1,3-butadiene 0.5

Hexachloroethane 3.0

Lead 5.0

Lindane 0.4

Mercury 0.2

Methoxychlor 10.0

Methyl ethyl ketone 200.0

Nitrobenzene 2.0

Pentachlorophenol 100.0

Pyridine 5.0

Selenium 1.0

Silver 5.0

Tetrachloroethylene 0.7

Toxaphene 0.5

Trichloroethylene 0.5

2,4,5-Trichlorophenol 400.0

2,4,6-Trichlorophenol 2.0

2,4,5-TP (Silvex) 1.0

Vinyl chloride 0.2

a If o-, m-, and p-cresol concentrations cannot be differentiated, the

total cresol concentration is used.

13
set national standards for the quality of ambient air and

the states must ensure that these standards are met.

States generally regulate sources of air contaminants on

a case-by-case basis. WTPs that use aeration technologies, for example, might be subject to state air quality

regulations because of concerns about human exposure

to air emissions.

The CAA, initially passed in 1970 and amended in 1977,

1988, and 1990, gives EPA authority to set national

standards for the quality of ambient air and to regulate

sources of pollution that may affect air quality. The cornerstone of the CAA is a set of National Ambient Air

Quality Standards (NAAQS) for six pollutants: ozone,

total suspended particulates, sulfur oxides, lead, nitrogen dioxide, and carbon monoxide. The NAAQS establish the maximum allowable concentration for each

pollutant in all areas of the United States (see Table 2-3).

The federal NAAQS do not specify the means by which

the emission levels are to be met. States must ensure

that NAAQS are achieved by enforcing all national emission standards and implementing any additional controls

necessary for their particular region. States must publish, and EPA must approve, State Implementation Plans

(SIPs) that describe the measures to be taken to ensure

that NAAQS are achieved and maintained.

Parts of the U.S. that fail to meet one or more of the

NAAQS are designated as nonattainment areas. Figure

2-1, for example, shows the location of nonattainment

areas for ozone. NAAQS have been established for

ozone, and VOCs are regulated by the states as ozone

precursors on a source-by-source basis.

In addition to setting NAAQS and limiting emissions to

achieve these standards, EPA has implemented other

programs for controlling airborne contaminants. Under

Section 111 of the CAA, EPA has the authority to establish New Source Performance Standards (NSPS) for

restricting emissions from new industrial facilities or facilities undergoing major modifications. Under the CAA,

EPA must set the NSPS control levels that reflect the

“degree of emission reduction achievable” through use

of the best available control technology (BACT) that has

been “adequately demonstrated.”

For toxic air pollutants not covered by the NAAQS or

NSPS, EPA promulgates National Emission Standards

for Hazardous Air Pollutants (NESHAPs). NESHAPs

address pollutants with more limited exposure but more

extreme health effects than pollutants controlled under

the other standards. Most NESHAPs are defined in

terms of the rate of emission from a source. Thus far,

NESHAPs have been promulgated for eight compounds: arsenic, asbestos, benzene, beryllium, mercury, polycyclic organic matter (POM), radionuclides,

and vinyl chloride. For example, NESHAPs governing

radionuclide emissions require that the maximum radiation dose to an individual be no more than 10 mrem/yr.

Facilities must monitor their emissions at doses of 1

percent of the limit or 0.1 mrem/yr.

2.8.2 State Regulations

As indicated earlier, NAAQS are established by EPA for

specific pollutants and the states then set standards to

attain and maintain them. Each state’s approach and

timetable for ensuring compliance with NAAQS are

summarized in its SIP. The SIP can incorporate many

regulatory measures that go beyond federal emissions

limitations to achieve compliance with the NAAQS. Each

SIP is reviewed and approved by EPA.

Individual states differ considerably in their approaches

and timetables for regulating air emissions. In general,

states do not have specific requirements for WTPs;

instead, they evaluate emissions of air contaminants on

Table 2-3. National Ambient Air Quality Standards (40 CFR

Part 50)a

Carbon monoxide Primary: 35.0 parts per million (ppm)

averaged over 1 hour and 9.0 ppm

averaged over 8 hours; neither level to be

exceeded more than once per year.

Secondary: Same as primary.

Fine particulate

matterb Primary: 150 μg/m3 averaged over 24

hours, with no more than one exceedance

per year averaged over a 3-year period;

also, 50 μg/m3 expected annual arithmetic

mean.

Secondary: Same as primary.

Lead Primary: 1.5 μg/m3 arithmetic average over

a quarter of a calendar year.

Secondary: Same as primary.

Nitrogen dioxide Primary: 100 μg/m3 (or 0.053 ppm) as

annual arithmetic mean concentration.

Secondary: Same as primary.

Ozone Primary: 235 μg/m3 (0.12 ppm) averaged

over 1 hour, not to be exceeded more than

once per year. (The standard is satisfied if

the number of calendar days on which the

standard is exceeded is 1 or less. Multiple

violations in a day count as one violation).

Secondary: Same as primary.

Sulfur oxides Primary: 365 μg/m3 (0.14 ppm) averaged

over a 24-hour period, not to be exceeded

on average more than once per year over

a 3-year period; 80 μg/m3 (0.03 ppm)

annual arithmetic mean.

Secondary: 1,300 μg/m3 average over a

3-hour period, not to be exceeded more

than once per year.

a National Primary and Secondary Ambient Air Quality Standards,

July 1, 1987. b Standard applies only to particulate matter that is ≤ 10 μm in

diameter.

14
a source-by-source basis. For example, some state

requirements, along with local air quality regulations,

limit gas phase emissions from stripping processes

(mechanical or packed tower) and from reactivation systems using GAC because of the possibility of generating

contaminants such as VOCs or radon. Some states do

not permit radionuclide air emissions, thereby preventing the use of the packed tower aeration process. Other

states require radon off-gas treatment. Some states only

require permit applications from source operations

where emissions exceed the maximum allowable rate

set by the state.

Michigan has an extensive air quality program that encompasses emissions from WTPs. The Michigan Air

Pollution Act authorizes the Michigan Air Pollution Control Commission (MAPCC) and the Michigan Department of Natural Resource’s Air Quality Division to issue

permits for the installation and operation of equipment

or processes that may emit air contaminants. Based on

this authority, MAPCC has promulgated a set of General

Rules (amended April 17, 1992) concerning air use approval. These regulations require the issuance of an air

use permit from MAPCC for the installation or modification of any process or equipment that may emit an air

contaminant.

After installing the equipment, MAPCC must issue an

operating permit before the equipment can go into operation full time. Rule 285 of MAPCC General Rules

outlines permit system exemptions. Subsection (j) of this

rule exempts lagoons and sewage treatment equipment

from permitting requirements except for lagoons and

equipment primarily designed to treat VOCs in wastewater or ground water, unless the emissions from these

lagoons and equipment are only released into the general in-plant environment.

Air use permit applications submitted to MAPCC must

include descriptions of the equipment; the site; the exhaust system configuration; data on the exhaust gas

flow rate; an operating schedule; and information on any

air pollutants to be discharged. Application review consists of a technical evaluation by the Permit Section

engineers and a site evaluation by the Compliance Section district staff. After internal processing is completed,

the Air Quality Division develops the necessary permit

conditions and stipulations to ensure that the proposed

Figure 2-1. Areas exceeding the ozone NAAQs (U.S. EPA, 1992a).

15
process or plant operates in an environmentally safe

and acceptable manner. Under Michigan law, failure to

obtain or comply with a permit can result in fines of up

to $10,000 and additional fines of up to $2,000 per day

for as long as the violation continues.

In California, local districts develop their own regulations

and permitting requirements for stationary sources of air

emissions. These districts regulate operations that result in air emissions on a source-by-source basis. The

Bay Area Air Quality Management District, for instance,

has no specific regulation governing air emissions from

WTPs but it does have permitting requirements for air

stripping processes. California’s South Coast Air Quality

Management District has a nuisance regulation (Rule

402) that prohibits discharging, from any source, air

contaminants that can cause injury, nuisance, or annoyance to the public. WTPs generating air emissions in this

district would be evaluated on a case-by-case basis.

Applications for air use must be submitted to the district.

The State of New Jersey has not promulgated regulations that specifically address air emissions from WTPs.

Subchapter 16 of New Jersey’s Air Pollution Control Act,

however, outlines requirements for the control and prohibition of air pollution by VOCs. Section 16.6(a) of the

law prohibits emitting VOCs into the atmosphere from

any source, in excess of the maximum allowable emission rate as set in the regulation. In New Jersey, WTPs

generating VOCs would be evaluated on a case-bycase basis.

16
Chapter 3

Characterization of Water Treatment Plant Residuals

The majority of residuals from water treatment plants

(WTPs) fall into one of four categories:

• Naturally occurring, colloidal/particulate matter (e.g.,

clay, silt, algae) removed from raw water by sedimentation, filtration, membranes, or other processes; and,

inert material in treatment chemicals (e.g., grit in lime).

• Naturally occurring, soluble substances (e.g., iron,

manganese, calcium, and magnesium) converted to

their insoluble precipitate forms by oxidation or pH

adjustment.

• Precipitates formed (e.g., Al(OH)3, Fe(OH)3) when

chemicals are added to water.

• Spent materials (e.g., granular activated carbon [GAC],

powdered activated carbon [PAC], filter media, resins) that must periodically be removed from unit treatment processes after exceeding their useful lives.

These residuals are addressed in this chapter, which

includes a discussion of sludges, liquid wastes, radioactive wastes, and physical and chemical characteristics.

3.1 Types and Quantities of Residuals

3.1.1 Sludges

Semi-solid residuals produced from mechanical water

clarification processes (e.g., screenings, presedimentation), as well as those produced from the clarification of

water that has been chemically preconditioned, are generally referred to as sludges. The three most common

types of sludges are coagulant/polymeric, chemical softening, and oxidized iron/manganese.

If a raw water source has a high concentration of total

suspended solids (TSS) the alum/iron coagulant

sludges will contain a high percentage of gelatinous,

hydroxide precipitates (e.g., Al(OH)3, Fe(OH)3), and will

exhibit the overall characteristics indicated in Table 3-1.

Iron/manganese sludges also tend to be composed of

gelatinous hydroxide solids (e.g., Fe(OH)3, Mn(OH)2).

Chemical softening sludges primarily consist of crystalline calcium carbonate (CaCO3), with the magnesium

hydroxide (Mg(OH)2) portion of the solids increasing as

the magnesium content of the raw water increases.

Because of the structured, crystalline nature of the

CaCO3 precipitate, the relationship between solids content and the overall sludge characteristics of chemical

softening sludges is different from that for coagulant and

iron/manganese sludges (see Table 3-2).

Several equations can be used to predict the quantity of

alum/iron coagulant sludge to be generated, based on

the raw water characteristics and amount of coagulant

dose. The principal factors used in the estimation of

coagulant sludge quantities are: 1) the suspended solids

(SS)-to-turbidity ratio for the raw water (Cornwell et al.,

1987); and 2) the waters of hydration assumed for the

coagulant. Perhaps the most commonly used equations

for predicting the quantity of alum or iron coagulant

sludge are (Cornwell et al., 1987):

S = (8.34Q)(0.44Al + SS + A) (Eq. 3-1)

where

S = sludge produced (lbs/day)

Q = plant flow, million gallons per day (mgd)

Al = liquid alum dose (mg/L, as 17.1% Al2O3)

SS = raw water suspended solids (mg/L)

A = net solids from additional chemicals added

such as polymer or PAC (mg/L)

Table 3-1. Alum/Iron Coagulant Sludge Characteristics

(ASCE/AWWA, 1990)

Solids Content Sludge Characteristic

0–5% Liquid

8–12% Spongy, semi-solid

18–25% Soft clay

40–50% Stiff clay

Table 3-2. Chemical Softening Sludge Characteristics

(ASCE/AWWA, 1990)

Solids Content Sludge Characteristic

0–10% Liquid

25–35% Viscous liquid

40–50% Semi-solid

60–70% Crumbly cake

17
and

S = (8.34Q)(2.9Fe + SS + A) (Eq. 3-2)

where

Fe = iron dose (mg/L, as Fe)

Similar equations can be used to predict the quantity of

sludge produced when calcium and magnesium, carbonate and noncarbonate hardness is chemically precipitated. The quantity of sludge produced depends on

whether lime/soda ash or caustic soda is used as the

softening chemical(s) and on the total amount of hardness that is removed. Obviously, much more sludge is

produced when lime, rather than caustic soda, is used

to precipitate carbonate hardness, since the calcium

associated with the lime must also be precipitated in the

chemical softening process.

The quantity of sludge produced when soluble iron (II)

and manganese (II) are oxidized to their insoluble

precipitate forms (i.e., Fe(III) and Mn(IV)) depends on

several factors. The factor that most affects the sludge

quantities is the oxidant used (e.g., oxygen, permanganate, chlorine dioxide, ozone). Similar to using lime to

precipitate calcium from hard water, using permanganate to oxidize iron or manganese results in more

sludge. The manganese associated with the permanganate is reduced from a (VII) to (IV), and is precipitated

along with the iron and/or manganese being oxidized.

3.1.2 Liquid Wastes

Perhaps the most common liquid waste generated at

WTPs in the past has been spent filter backwash water.

The spent filter water associated with filter-to-waste (rewash) has become more common as WTPs prepare for

compliance with the Surface Water Treatment Rule.

Slow sand filter wastes are also becoming more prevalent as some smaller communities return to the combined physical/biological benefits of slow sand filtration

as described in Section 3.1.2.2. Regenerate wastes (i.e.,

brine and rinse water wastes) associated with ion exchange (IX) facilities continue to be produced by some

softening plants. Reject waters from various membrane

processes (e.g., reverse osmosis, nanofiltration, ultrafiltration, microfiltration) are gaining prominence as maximum contaminant levels for finished water are set at lower

levels for more organic and inorganic contaminants.

3.1.2.1 Spent Filter Backwash Waters

Spent filter backwash water generally represents a volume of 2 to 5 percent of the total water processed at a

WTP. The suspended solids concentration of spent filter

backwash varies throughout the 10- to 15-minute duration of the backwash, with the water gradually becoming

cleaner as the backwash proceeds. The average suspended solids concentration of spent backwash typically

falls within the range of 50 to 400 mg/L.

Filter backwash water historically has been returned to

the head of a WTP to be processed again. An equalization basin is usually used so that the spent backwash

water can be returned to the head of the WTP at a rate

less than 10 percent of the raw water flow into the WTP.

Concerns over the recycling of microorganisms, aggravation of taste and odor problems, increase in disinfection byproducts, and other issues have drastically

reduced the number of WTPs that directly recycle spent

filter backwash. Thus, in recent years interest has been

generated in better understanding the characteristics of

backwash wastes, since these must be processed along

with other WTP residual streams.

Another form of waste from filters that is becoming

popular again is filter-to-waste, or rewash, which refers

to the wasting of filtered water during the ripening stage

of a clean filter. Concern over the passage of certain

microorganisms (e.g., Giardia, Cryptosporidium, and viruses) through the filter media of a freshly backwashed

filter has renewed interest in filter-to-waste.

The filter-to-waste period for ripening a freshly backwashed filter at most WTPs ranges from 15 minutes to

an hour in length. Some WTPs are finding that the length

of this filter ripening period can be shortened by introducing a coagulant aid or by allowing a filter to sit idle

for a time between when it has been backwashed and

is returned to service. The filtration rate used to ripen a

filter varies from one WTP to another. While many WTPs

filter-to-waste at the normal filtration rate, some plants

filter at only a fraction of the normal rate. Often, this

variation is due to the fact that a smaller pipe is available

to convey the filter-to-waste flow. The filter-to-waste

flow, while not considered to be of a quality that it can

be sent directly into the distribution system, is generally

a fairly clean waste stream. Therefore, at most WTPs

this flow is equalized and returned to the head end of

the plant. Other options being used to handle the filterto-waste flow include discharging it to a local storm

sewer with an appropriate National Pollutant Discharge

Elimination System (NPDES) permit; discharging it to a

sanitary sewer for processing at a local wastewater

treatment plant (WWTP); introducing it to the solids

handling stream of the WTP; or, treating it with a membrane process prior to returning the flow to the head end

of the plant.

3.1.2.2 Slow Sand Filter Wastes

Slow sand filtration is a simple, economical, and generally reliable method of treating low turbidity waters for

potable uses. Since application rates for slow sand filtration are low, on the order of 40 to 150 gallons per day

per square feet (gpd/ft2

), the process is mostly used by

smaller treatment facilities. Organics, silt, and other particles are trapped in the upper portions of the filter, which

is periodically removed for cleaning. This process is called

18
scraping and is normally performed by hand, but small

mechanical equipment is sometimes used. The filter is

resanded after several scrapings when the filter depth

reaches a predetermined minimum design thickness.

Material for resanding can be either new sand or old

filter sand that has been removed and washed. A unique

feature of the slow sand filter is the schmutzdecke, a

biologically active layer in the top of the filter. Viruses,

cysts, and other organisms are almost entirely removed

during slow sand filtration. Thus, the sand removed

during scraping can contain a fairly active biological

population.

Sources of Liquid Residuals

• Scraping: Normal scraping removes the top 0.5 to 1

inch of sand (Tanner and Ongerth, 1990), although

scraping to a depth of 4 inches has been reported

(U.S. EPA, 1985a). Scraping generally is conducted

at a given headloss or reduction in flow, but may be

initiated on a regular cycle without regard for headloss. Consequently, scraping may occur from three

to forty times annually. Some facilities dispose of the

removed material by stockpiling it for other uses such

as road sanding during the winter or as soil additives.

More commonly, the material is washed and then

stored for later addition back to the filter. In this case,

the wash water constitutes a residuals stream that

may require treatment. Current common disposal

methods include discharging to a sewer or a receiving watercourse without treatment. Discharge to a

receiving watercourse may require a state or EPA

regional NPDES permit.

• Raking: Slow sand filters are sometimes raked, usually by hand with a garden rake, to loosen the top

layer of material and improve the hydraulic rate without removing sand. This process normally does not

produce any waste residuals. Wet-harrow cleaning is

sometimes used; this procedure uses a flow of water

to flush the raked deposits from the filter, and these

deposits may require treatment (Logsdon, 1991).

Facilities that practice raking normally need to remove more sand during scraping than facilities that

do not rake.

• Backwash water: One facility reportedly reverses the

flow through the slow sand filter without expanding

the bed. The filter is raked prior to backwashing,

which promotes removal of the schmutzdecke (Tanner and Ongerth, 1990). The backwash water is normally discharged directly to the receiving water, but

may require treatment at some locations.

• Filter-to-waste: A fairly large volume of wastewater

can be generated during filter-to-waste cycles as

some WTPs waste the filtered water for 24 to 48

hours before slow sand filters are placed back on line

after cleaning. A filter-to-waste procedure is specified

under the current edition of “Ten State Standards”

(Great Lakes, 1992). The high quality of this water

normally allows disposal without treatment.

Quantities of Residuals

• Scraping: Scraping can remove sand to depths of 0.5

to 4 inches. Removing 1 inch of sand will generate

2 to 6 ft3

 of material per 1,000 gpd of filter design

capacity, based on design rates of 45 to 150 gpd/ft2

.

• Spent backwash water: Because slow sand filters are

seldom subjected to any type of backwashing, the

majority of WTPs do not generate this particular residuals stream. A slow sand filter in northern Idaho

was periodically cleaned by backwashing and scraping, and frequently exhibited higher-than-desired filtered water turbidities (Tanner and Ongerth, 1990).

The extent, if any, to which backwashing contributed

to this problem is not known, but backwashing is not

normally a recommended practice.

• Filter-to-waste: Filtering to waste after cleaning is a

recommended operating practice, and is required by

some states. High quality filtered water is normally

discharged to waste without treatment. At some locations, however, this residuals stream may be subject to provisions of an NPDES permit. Waste

volumes are generally in the range of 200 to 600

gal/hr/100 ft2

 of slow sand filter area. Filtering-towaste periods are normally of 24- to 48-hour duration

but vary from site to site.

3.1.2.3 Ion Exchange Brine

IX offers the possibility of removing one or more ionic

species from one liquid phase and transferring them to

another liquid phase via an intermediate solid. In many

cases the transfer can be made on a selective basis and

with good chemical efficiency. The IX process is primarily used to remove hardness from water, particularly in

small water systems and for residential use. IX also has

been shown to be effective in removing nitrates and

other contaminant ions, including barium, radium, arsenate, selenate, fluoride, lead, and chromate (Pontius,

1990).

In the application of IX to remove hardness, the hardness in the water (most often Ca2+ and Mg2+) exchanges

with an ion from the resin (generally sodium, Na+) because the resin prefers the contaminant ions. The reactions are as follows (where X represents the solid IX

material):

Carbonate Hardness

Ca(HCO3)2 + Na2X → CaX + 2NaHCO3

Mg(HCO3)2 + Na2X → MgX + 2NaHCO3

19
Noncarbonate Hardness

CaSO4 + Na2X → CaX + Na2SO4

CaCl2 + Na2X → CaX + 2NaCl

MgSO4 + Na2X → MgX + Na2SO4

MgCl2 + Na2X → MgX + 2NaCl

Regeneration

CaX + 2NaCl → CaCl2 + Na2X

MgX + 2NaCl → MgCl2 + Na2X

Through these reactions, calcium, magnesium, and

other materials are removed from the water and replaced by an equivalent amount of sodium (i.e., two

sodium ions for each divalent cation removed).

The exchange results in nearly 100 percent removal of

hardness from the water; in the process, the resin becomes saturated and reaches its exchange capacity. At

this point, breakthrough occurs and hardness can no

longer be completely removed from the water. In practice, there is competition for the IX sites by elements

other than calcium or magnesium; these other constituents may also limit the effectiveness of the resin at

removing hardness. When this situation occurs or the

resin is saturated, the IX material is regenerated. Regenerant water containing a large excess of Na+ (such

as a concentrated NaCl solution) is passed through the

column to remove the hardness. The mass action of

having a large excess of Na+ in the water causes ions

on the resin to be replaced by Na+ and then enter the

water phase. This constitutes a reversal of the initial IX

reactions. The regenerated IX material can then be used

to remove more hardness (Cornwell et al., 1987).

In addition to this regenerant waste—consisting of sodium, chloride, and hardness ions—wastes produced by

IX processes include the backwash water and rinse

water used, respectively, before and after the formal

regeneration of the resin. The term regenerant waste, or

brine, is frequently applied to the combination of the

used regenerant and the slow rinse, the initial portion of

the rinse. When comparing data, it is important to know

whether concentrations are reported as only concentration in the regenerant waste itself, or as diluted with rinse

water and/or backwash water.

Solids Content

Spent brine often has a very high concentration of total

solids and total dissolved solids (TDS). Brine wastes

usually contain very few suspended solids. Gradually,

resins lose their capacity to be regenerated and upon

being replaced, become a solid waste.

pH

IX brine from the water softening process is generally of

neutral range pH. More generally, the pH of the brine

depends on the nature of the regenerant. Other cation

exchange resins may be regenerated using concentrated sulfuric or hydrochloric acid. The resulting regenerant waste will need to be neutralized from its low pH,

which is primarily determined by how much excess acid

was present in the regenerant solution. Anion exchangers are usually regenerated with a basic material or

sodium chloride. Weak basic resins (which will remove

strong anions such as chloride, sulfate, and nitrate) are

typically regenerated using sodium carbonate. Strong

basic resins (which will remove most anions such as

chlorides, sulfate, nitrate, bicarbonate, and silica) are

regenerated using sodium hydroxide. In this case, the

high pH of the resultant regenerant waste will need to

be neutralized before being discharged.

Ion Content

Regenerant waste contains the hardness removed from

the resin, chloride from the regenerant solution, sodium

present as excess regenerant, and smaller amounts of

other ions removed from the IX resins. Table 3-3 shows

the typical ranges of ion concentrations in the wastewater.

The chemical concentration of brines varies widely from

plant to plant, depending on raw water hardness (concentration of the cations to be removed), regenerant

dose and concentration, rinsing procedures, and cation

exchange capacity of the resin.

Because the exchange of hardness for sodium ions is

stoichiometric—that is, one equivalent for one equivalent—the total equivalents of ions in the wastewater will

equal the original number of equivalents in the regenerant. The distribution of these equivalents between the

various chemical species depends on the IX capacity of

the resin and its regeneration efficiency, or taken together, the operating capacity of the resins. The operating capacity is always less than advertised exchange

capacity because of incomplete regeneration and contaminant leakage (breakthrough of the least binding

counter-ion prior to saturation of the resin).

Table 3-3. Typical Chemical Constituents of Ion Exchange

Wastewater (AWWARF, 1969)

Constituents Range of Averages (mg/L)

TDS 15,000–35,000

Ca++ 3,000–6,000

Mg++ 1,000–2,000

Hardness (as CaCO3) 11,600–23,000

Na+ 2,000–5,000

Cl– 9,000–22,000

20
Toxicity

IX brine typically has a high TDS content, with the

predominant ion being the co-ion in the regenerant

(chloride, sulfate, carbonate, or sodium, depending on

the type of resin and, consequently, the type of regenerant used). The other ions present are those removed

by the resin and reclaimed by the regenerant. With

chlorides present, brine is corrosive to materials it contacts. Brine possesses varying levels of toxicity to the

environment, depending on its TDS level and specific

chemical makeup.

Quantity

The total amount of wastewater (spent brine) usually

ranges from 1.5 to 10 percent of the amount of water

softened, depending on the raw water hardness and the

operation of the IX unit (AWWARF, 1969; O’Connor and

Novak, 1978).

Figure 3-1 shows the expected wastewater volume as a

function of raw water hardness for the case where all

other variables are held constant. Table 3-4 reflects how

the parameters mentioned above influence the quantity

of brine wastewater produced.

3.1.2.4 Reverse Osmosis, Nanofiltration, and

Electrodialysis-Electrodialysis Reversal

Process Description

Reverse osmosis (RO), nanofiltration (NF) (or membrane softening), ultrafiltration (UF), and microfiltration

(MF) membrane processes use semipermeable membranes to remove contaminants from a feedwater. Typically pretreated to minimize scaling and fouling of the

membrane, feedwater flows across the membrane surface. Increase in pressure on the feedside of the membrane transports some of the water from the feedside,

through the membrane, to the permeate side. Other

constituents may be rejected by the membrane, to an

extent that depends on the properties of the particular

membrane used and the operating conditions. Water

and other constituents not permeating the membrane

flow out of the membrane system as the concentrate

stream. Table 3-5 provides a system operation summary

for these membrane processes. Each process has different operating conditions and ranges of feedwater TDS.

At one extreme of high rejection RO membranes, typically used for seawater or brackish water applications,

ions may be almost completely rejected (rejections

greater than 99 percent). The larger, more effective

pore-sized NF membranes reject ions to a lesser degree. Here, the monovalent ions may be rejected to

perhaps 50 to 70 percent and multivalent ions (hardness) to perhaps 90 percent levels. Other membrane

processes have lower rejection rates of dissolved ions.

Species other than dissolved ions, such as dissolved

organics, dissolved gases, biological contaminants, and

suspended solids, not removed in pretreatment steps,

750


650


550


450


350


250


150


50

10,000 20,000 30,000 40,000 50,000 60,000 70,000 80,000 

Total Wastewater Volume, gal/MG Treated

Total Hardness, mg/L

Figure 3-1. Generation of wastewater volumes with ion exchange (U.S. EPA, 1984a).

Table 3-4. Regeneration of Cation Exchange Resins (AWWARF, 1969)

Plant

Gal

Wastewater/

1,000 Gal

Water

Processed

Raw Water

Total

Hardness

(mg/L as

CaCO3)

Gal

Regenerant/ft3

Resin

Gal

Rinse/ft3

Resin

Concentration

of Brine

(lb/gal)

Gal

Processed/ft3

Resin-Cycle

Dosage 

(lb salt/ft3

resin) Reference

Crystal Lake Plant #6 21.9 233 7.3 19.4 0.90 1,220 6.6 U.S. EPA, 1984a

Crystal Lake Plant #8 17.2 244 5.1 19.0 1.26 1,400 6.5 U.S. EPA, 1984a

Eldon 71.9 375 3.9 61.7 1.43 750 5.6 U.S. EPA, 1976a

Grinnell 49.5 388 14.5 35.0 0.50 1,000 7.2 U.S. EPA, 1976b

Holstein 53.5 885 5.7 19.7 1.16 475 6.6 U.S. EPA, 1976b

Estherville 82.8 915 4.4 24.7 1.25 350 5.5 U.S. EPA, 1976b

Note: The chemical characteristics of the backwash water did not show a large variation from the raw water; therefore, they were not included

in this analysis.

21
are similarly rejected according to their size and interactions with the membrane.

In the electrodialysis (ED) process there is no pressure

used and, thus, no significant bulk flow of water through

the membrane. The process employs sets of anion- and

cation-selective membranes, electrodes placed outside

of the sets of membrane pairs, and an impressed electrical potential across the entire membrane stack. Several flow channels are created between the membrane

pairs and between the outer membranes and the electrodes. Water is fed to all channels. The electrical force

causes movement of the cations and anions in different

directions and out of certain channels into other channels. Some channels become relatively depleted in ions

and others become more concentrated. All product

streams are combined and all concentrate streams are

combined upon exit from the membrane stack. Electrodialysis reversal (EDR) is an ED process in which the

polarity of the electrodes is reversed on a prescribed

time cycle, thus reversing the direction of ion movement

in a membrane stack. The effective rejection of dissolved ions in the ED-EDR process can be quite high

(above 90 percent). The EDR process has several advantages over the straight ED system, including lower

membrane fouling, less need for pretreatment, and

higher reliability. Virtually all new ED plants are of the

EDR type. ED-EDR processes do not remove nonpolar

contaminants.

UF and MF membrane systems are receiving increased

attention because of their potential to remove particulates, microorganisms, and larger organics—constituents addressed in the Safe Drinking Water Act (SDWA)

amendments. This discussion, however, is restricted to

the RO, NF, and ED-EDR membrane processes.

The recovery rate of these membrane systems refers to

the percent of feedwater that is converted into product

water. Thus, a 60 percent recovery process would have

a concentrate stream of 40 percent of the feedwater flow.

Table 3-6 shows the applications for the RO, NF, and

ED-EDR processes and the removal characteristics of

the membrane processes. For producing drinking water,

NF is limited to freshwater (lower TDS) applications and

ED-EDR to applications calling for removal of polar

constituents in a nonselective manner.

As of September 1992, 137 drinking water membrane

plants sized 25,000 gpd or greater, of the various types

mentioned, existed in the continental United States

(Mickley et al., 1993). Of these plants, 101 (74 percent)

were brackish water RO plants, 16 (12 percent) were NF

plants, 13 (9 percent) were ED-EDR plants, and the

remaining 7 (5 percent) were seawater RO plants.

Nature of the Concentrate

Most residuals, or more generally, industrial wastes, are

characterized by the chemicals added during the processing of the waste. Membrane concentrate, on the

other hand, has very few process-added chemicals and

thus reflects the character of the raw water used. Membrane processes do not produce more pollutant material

or mass—they redistribute, or concentrate, those constituents present in the raw water that are rejected by

the membrane.

Generally, posttreatment concentrate contains all the

species in the raw water that are not removed in pretreatment. Pretreatment typically consists of acid addition, antiscalant addition, and a 5-μm filtration step.

Species that are not rejected by the membrane are

present in both the product and concentrate at the same

concentration as in the original feedwater. Species that

are rejected by the membrane are concentrated to an

extent depending on the membrane rejection of the

particular species and the membrane system recovery

(defined as the percentage of the feed flow recovered

as permeate product). The nature of the concentrate

may be affected by posttreatment.

Table 3-5. Membrane Process Operations Summary (Mickley et al., 1993)

Membrane Process Feedwater TDS (mg/L)

Typical

Operating

Pressure (psi)

Typical System

Recovery (%) System Rejection (%)

Seawater RO 10,000–45,000 (high P) 800–1,200 20–50 99+ (TDS)

Brackish RO 500–3,500 (low P)

3,500–10,000 (medium P)

100–600 60–85 85–96 (TDS)

95–98 (hardness)

Nanofiltration Up to 500 50–150 75–90 80–90 (hardness)

Ultrafiltration,

microfiltration

Not used to remove TDS Below 100 90 and above Zero rejection of TDS; rejection

of other species dependent on

specific membrane

Electrodialysis (ED) Up to 7,500 (not economical

at high TDS)

Not applicable 70–90+ (Effective monovalent removal

can be >95 for ED and >99+ for

EDR)

Key

P = phosphorus.

TDS = total dissolved solids.

22
Because membrane concentrate is defined by the raw

water characteristics and raw water characteristics are

site specific, the specific nature of the concentrate is

also site specific.

Solids

Due to the need to prevent plugging of membrane system flow channels, allowable feedwater levels of suspended solids are fairly low for the RO, NF, and ED-EDR

systems. Depending on the specific flow channel dimensions, the feedwater limits may range from 1 to 5 SDI

units. In general, therefore, the levels of suspended

solids are low in the feedwater, and these levels are

concentrated to a degree that depends on the system

recovery.

The TDS level can vary over a wide range, depending

on the TDS level of the feedwater and the membrane

system rejection and recovery. Table 3-7 provides

ranges for feedwater TDS for the various membrane

processes. The amount of TDS in the concentrate depends on the rejection level of the particular membrane

system and the process recovery.

pH

The pH of the concentrate depends on the pretreatment

pH adjustments and possible posttreatment pH adjustments. The pretreatment adjustments are dictated by

the need to control the scaling potential from solubilitylimited species, such as calcium carbonate, and by the

pH limits of the particular membrane. Typically, the feedwater is acidified to a range of about 5.5 to 7.0. The pH

of the concentrate is generally higher than the feedwater

pH, due to distribution of the carbonate species between

the concentrate and the product streams. Posttreatment

pH adjustment may be used to render the concentrate

compatible with the receiving water, in the case of disposal to surface waters. Adjustment may be used to ensure the noncorrosive nature of the concentrate in cases

where the discharge will be exposed to piping or confining vessels prior to, or as part of, the disposal method.

Quantity

The quantity of concentrate is directly related to the

recovery, R, of the membrane system. Equation 3-3 can

be used to calculate the quantity of concentrate that is

generated by the membrane system.

Qc = Qf(1 − R) (Eq. 3-3)

where

Qc = quantity of concentrate flow

Qf = quantity of feedwater flow

R = recovery rate of the membrane system

Qc may also be expressed in relation to the product

volume flow (Equation 3-4).

Qc = Qp(1 − R)/R (Eq. 3-4)

where

Qp = quantity of product volume flow

R = recovery rate of the membrane system

For example, if the feed flow is 2 million gallons per day

(mgd) and the recovery is 70 percent (R = 0.70), then

Qp equals 1.4 mgd and Qc equals 0.6 mgd from either

of the above expressions.

Table 3-6. Membrane Process Applications for RO, NF, and ED-EDR (Mickley et al., 1993)

Applications Reverse Osmosis (RO) Nanofiltration (NF)

Electrodialysis–Electrodialysis

Reversal (ED-EDR)

TDS reduction

Seawater desalting Very effective Not effective No (not economical at high TDS)

Brackish water desalting Very effective Dependent on feedwater

makeup

Very effective

Freshwater treatment

Hardness ions removal (softening) Very effective Very effective No

Dissolved organics removal Very effective Very effective No

Color removal Very effective Very effective No

THM precursor reduction Very effective Very effective No

Specific inorganic and radionuclide

removal

Very effective Dependent on feedwater

makeup

Dependent on feedwater 

makeup

Key

TDS = total dissolved solids.

THM = trihalomethane.

Table 3-7. Concentration Factors for Different Membrane

System Recoveries (Mickley et al., 1993)

Recovery (%) Concentration Factor

50 2.0 

60 2.5 

70 3.33

80 5.0 

90 10.0 

23
The recovery of a membrane process is generally limited

by the potential for sparingly soluble salts, as they become

concentrated, to precipitate and scale membranes.

Toxicity

The potential toxicity concerns for concentrate have to

do with aquatic species, in the case of surface discharge, and with vegetation, in the case of land applications such as spray irrigation. Potential sources of

toxicity include individual components in the raw water

that are concentrated in the membrane process to an

extent that they are toxic to the life form in question. In

general, drinking water membrane concentrate is not

toxic as long as the affected life forms are matched to

the general TDS level of the concentrate. In the incidences where concentrate is potentially toxic, dilution of

the concentrate is generally a means of rendering the

concentrate nontoxic.

Ionic Content and Prediction of Concentrate

Concentrations

For species that are completely rejected by the membrane and thus totally retained in the concentrate, the

degree of concentration, or the concentration factor (CF)

may be defined as:

CF = 1/(1 − R) (Eq. 3-5)

where

R = the fractional system recovery

The CFs for different recoveries are shown in Table 3-7.

For example, if the feedwater has a TDS level of 10,000

parts per million (ppm), the rejection of the membrane

is assumed to be complete, and the recovery is 60

percent, then the concentrate would have a TDS level

of 25,000 ppm (from 10,000 multiplied by a concentration factor of 2.5). Each constituent of the TDS would

similarly be present in the concentrate at 2.5 times the

feedwater concentration.

As suggested by Table 3-5, complete rejection of constituents is not always the case. For species that are not

completely rejected, concentration still takes place, but

to a lesser extent. A theoretical expression for this situation (Saltonstall and Lawrence, 1982) is

CF = 1/(1 − R)r (Eq. 3-6)

where

r = the fractional rejection for the species in question

Note that for the case where r = 1, Equation 3-6 reduces

to Equation 3-5.

Table 3-8 shows concentration factors calculated from

Equation 3-6.

Each constituent has its own characteristic rejection,

which is a function of the particular membrane, the

particular species, and the operating conditions. The

data in Table 3-8 provide some insight into the complexity of estimating concentrate concentrations, but do not

provide a simple means for doing this.

Estimates for concentrate concentrations require the

specification of raw water characteristics, the particular

membrane used, the membrane system configuration,

and the operating conditions. Membrane manufacturers

have computer programs that can provide design quality

estimates for this. These programs, however, do not

provide 1) a rapid and simple means of estimating, or 2)

a means for estimating the concentrations of constituents present in minor amounts—constituents that may

influence the permitting of concentrate disposal.

Analysis has shown that assumption of 100 percent

rejection is quite accurate for seawater and brackish RO

membranes (Mickley et al., 1993). Use of the 100 percent rejection assumption will be conservative in the

sense of overestimating the concentration of the concentrate, a worst-case scenario. This approach may

also be used for EDR systems. The errors should be in

the range of 15 percent and under. For NF membrane

systems, assumption of complete rejection of all species

leads to more significant errors. Consequently, it is recommended that rejections of 70 and 90 percent be used,

respectively, for monovalent and multivalent ions.

Other guidelines to be used in estimating concentrate

concentrations are:

• Heavy metals should be considered to be rejected as

multivalent ions.

• Organics cover a wide range of molecular weights,

from less than 100 to greater than 100,000. Most are

rejected to a high degree in RO systems, somewhat

less so for NF systems, and to a low degree in EDR

systems (unless the organics in question are polar in

nature). For RO and NF systems, organics of molecular weight of approximately 1,000 may be considered to be completely rejected.

• Non-ionized gases have a rejection of zero in these

systems. Thus, the product and concentrate concentrations are the same as for the feed concentration.

• In RO and NF systems, the pH of the permeate will

be lower than that of the feedwater, and the pH of

the concentrate will be greater than that of the feedwater. The changes are caused by the redistribution

of carbonate species across the membrane. The

amount of the pH change in permeate and concentrate depends on the feed pH, the recovery, and the

amount of carbonate species present. Typical feed

pH values are in the range of 5 to 7, depending on

the type of membrane. Concentrate pH values may

be up to 1 or 1.5 pH units higher than the feed pH,

24
prior to any pH adjustment that may be done in posttreatment.

• Although dependent on pH, any chlorine present as

essentially non-ionized HClO may be considered as

not rejected by membranes.

• Radionuclides should be treated as multivalent ions.

• Fluorine and bicarbonate have pH-dependent rejections that range from values of zero at pH of 5 to

typical monovalent rejections at pH of 7.

• Silica rejections are dependent on the type of membrane;

the rejections can range from 75 to 98 percent for RO

membranes to 20 to 70 percent for NF membranes.

3.1.3 Radioactive Wastes

The types and quantities of radionuclides in residuals

depend on the ability of the WTP to remove specific

radionuclides from the drinking water. For example, under normal operation, cation exchange treatment will

remove radium, but not uranium or radon. Consequently, the regenerant brine waste from a cation exchange system will contain only radium, even though

the raw water may contain uranium and radon. Table 3-9

lists drinking water treatment processes, the radioactive

contaminant that they remove, and the types of residuals. This list is based upon the known ability of the water

treatment processes to remove the specific radionuclides either currently regulated or proposed for regulation for drinking water by EPA (U.S. EPA, 1991b). The

radionuclides radium-226, radium-228, uranium, and radon-222, are all naturally occurring.

Radon is a gas than can be removed from drinking water

by air stripping and GAC, neither of which produces a

residual for routine disposal. Furthermore, radon has a

very short half-life of approximately 3.5 days and decays

to essentially zero in roughly 28 days. Therefore, radon

should not be found in any waste stream from a conventional water treatment process, except in the air from an

air stripper. Lead-210, the next long-lived daughter product following radon-222 in the decay series, will be found

on any material that adsorbs radon, however.

Some materials used in drinking water treatment processes, either for direct removal of a contaminant such

as GAC and IX resins, or indirect removal of a contaminant such as filter sand in conventional treatment, will

adsorb radionuclides. When the time arrives for these

materials to be replaced, they will contain the radionuclides adsorbed but not removed from the material during the treatment process. A list of drinking water

treatment process materials and the potential radionuclides contained on these materials is provided in

Table 3-10.

The concentration of radionuclides in waste streams

produced by any water treatment process depends on

a number of factors: concentration of the radionuclide in

the source water, the percent removal of the contaminant, the volume of the waste streams, and the mode of

operation of the treatment process. These factors will

Table 3-8. Tabulation of Concentration Factorsa (Mickley et al., 1993)

Rejection Values

Recovery Values 1.0 0.99 0.98 0.95 0.90 0.80 0.70

0.90 10.0 9.77 9.55 8.91 7.94 6.31 5.61

0.88 8.33 8.16 7.99 7.50 6.74 5.45 4.41

0.86 7.14 7.00 6.87 6.47 5.87 4.82 3.96

0.84 6.25 6.13 6.03 5.70 5.20 4.33 3.61

0.82 5.56 5.46 5.37 5.10 4.68 3.94 3.32

0.80 5.00 4.92 4.84 4.61 4.26 3.62 3.09

0.78 4.55 4.48 4.41 4.21 3.91 3.36 2.89

0.76 4.17 4.11 4.05 3.88 3.61 3.13 2.72

0.74 3.85 3.79 3.74 3.60 3.36 2.94 2.57

0.72 3.57 3.53 3.48 3.35 3.14 2.77 2.44

0.70 3.33 3.29 3.25 3.14 2.96 2.62 2.32

0.68 3.13 3.09 3.05 2.95 2.79 2.49 2.22

0.66 2.94 2.91 2.88 2.79 2.64 2.37 3.13

0.64 2.78 2.75 2.72 2.64 2.51 2.26 2.04

0.62 2.63 2.61 2.58 2.51 2.39 2.17 1.97

0.60 2.50 2.48 2.45 2.39 2.28 2.08 1.90

a Calculated using the equation CF = 1 ÷ (1 - R)r

.

25
not be identical for any two plants and, consequently,

the concentration of radionuclides in the waste streams

will be site specific.

A limited amount of information has been reported on

the concentration of radionuclides in the waste streams

of several water treatment processes. Although these

data are site specific, these field measurements can be

used for approximating ranges, or levels, of radionuclide

concentrations that may be expected in these waste

streams.

3.1.3.1 Coagulation/Filtration Wastes

Laboratory, pilot plant, and full-scale system studies

have shown that uranium can be removed from source

water by conventional coagulation/filtration treatment,

and that removals are pH dependent (White and Bondietti, 1983; Lee and Bondietti, 1983; U.S. EPA, 1987).

Removals can range from 50 to 85 percent. The technical literature does not reveal any information on the

concentration of uranium in the waste streams of the

full-scale plants reported upon.

3.1.3.2 Lime Softening Wastes

Lime softening has been found to be very effective in

removing both radium and uranium, achieving removals

of up to 99 percent of both radionuclides (Lee and

Bondietti, 1983; Clifford, 1990; Bennett, 1978; Brink et

al., 1978; U.S. EPA, 1976a,b; Meyers et al., 1985; Sorg

and Logsdon, 1980; Jelinek and Sorg, 1988; Sorg,

1988). Removals are pH-dependent with highest removals achieved at pH levels of 10.0 to 10.5.

The lime softening process generates a lime sludge that

is precipitated during the process and which will contain

most of the uranium and radium removed during the

treatment cycle. A liquid waste is also produced by the

backwashing of the filter media. The backwash water

may be recycled to the front of the treatment plant, or

disposed of separately or with the lime sludge.

Field data showing specific measurements of radium226 and 228 in grab samples from full-scale WTP waste

streams are listed in Table 3-11. These data show a wide

range of radium concentrations in the waste streams.

For example, wet sludge from clarifier systems has

radium concentrations ranging from 980 to 4,577 pCi/L.

Dry weight concentrations vary from 2.8 to 21.6 pCi/g

for the same source of sludge. As expected, concentrations of radium-226 in lagoon sludges are higher with

volumes ranging from 5,159 to 11,686 pCi/L. Radium226 in filter backwash water ranges from 6 to 92 pCi/L.

No information was found in the literature on the concentration of uranium in lime softening treatment wastes.

3.1.3.3 Cation Exchange Wastes

The chemistry of radium is similar to that of calcium and

magnesium (hardness ions). Thus, cation exchange

resins in the sodium form used to soften water are very

capable of removing radium-226 and radium-228 from

drinking water (Clifford, 1990; Bennett, 1978; Brink et

al., 1978; U.S. EPA 1976a,b; Meyers et al., 1985; Sorg

and Logsdon, 1980). The cation IX regeneration process

produces three waste streams: backwash water, regenerant brine, and final rinse water. Although the regenerant brine contains most of the radium released from the

resin during the regeneration process, both the initial

backwash water and final rinse water will contain some

quantity of radium.

The concentration of radium in the waste streams is site

specific and depends on the method of plant operation

Table 3-9. Summary of Treatment Processes and the Types

of Wastes Produced From the Removal of

Radionuclides From Drinking Water

Treatment Process

Radionuclide

Removed

Types of

Residual/Waste

Coagulation/Filtration Uranium Sludge (alum/iron) 

Filter backwash water

Lime softening Radium, uranium Lime sludge

Filter backwash water

Cation exchange Radium Brine waste 

Backwash water

Anion exchange Uranium Brine waste 

Backwash water

Iron removal processes 

• Oxidation/Filtration

• Greensand adsorption

Radium Filter backwash water

Reverse osmosis Radium, uranium Reject water

Electrodialysis Radium, uranium Reject water

Air stripping Radon Airborne radon

Table 3-10. Water Treatment Process Materials Containing

Radionuclides

Treatment Process

Radionuclide

Removed Process Materials

Coagulation/Filtration Radium, uranium Filter medium (sand)

Filter medium (coal)

Lime softening Radium, uranium Filter medium (sand)

Filter medium (coal)

Cation exchange Radium Resin

Anion exchange Uranium Resin

Iron removal processes

• Oxidation/Filtration

• Greensand adsorption

Radium Filter medium (sand)

Filter medium (coal)

Greensand

Reverse osmosis Radium, uranium Membrane

Electrodialysis Radium, uranium Membrane

GAC adsorption Radon, uranium,

radium

GAC

Selective sorbents Radium, uranium Selective sorbent media

Key

GAC = granular activated carbon.

26
and regeneration employed. Examples of radium-226

concentrations in waste streams from several IX plants

are shown in Table 3-12. The average concentration of

radium-226 in backwash water, brine, and rinse water

together ranged from 22 to 82 pCi/L. The peak concentrations of radium-226 that occur during the brine cycle

ranged from 158 to 3,500 pCi/L.

3.1.3.4 Anion Exchange Wastes

Laboratory, pilot plant, and full-scale system studies have

shown that anion exchange treatment (chloride form) is

very effective for removing uranium from drinking water

(U.S. EPA, 1987; Jelinek and Sorg, 1988; Sorg, 1988).

This treatment process is similar to cation exchange for

radium and hardness removal. Anion resins are used in

the chloride form and are regenerated with sodium chloride (salt). Similar to cation exchange, the anion exchange process produces three waste streams during

regeneration: backwash water, regenerant brine, and

rinse water. Although the waste is a continuous stream,

each portion contains significantly different concentrations of uranium. The highest concentrations occur during brine regeneration; lower concentrations occur during

the initial backwash and rinse cycles.

The quantity and concentration of uranium in the waste

stream are site specific and depend upon the method of

plant operation and regeneration used. Most conventional anion resins have such high uranium removal

capacities that treatment runs could last as long as a

year or more. Current practice shows, however, that

regeneration generally is conducted once every month

or two to produce waste streams with lower uranium

concentrations. The concentration of uranium in the

waste streams depends to a great extent on the raw

water concentration and the length of the treatment run

(i.e., the volume of water treated).

Field data from a uranium removal plant treating water with

a uranium concentration of approximately 130 μg/L for two

months show peak uranium concentrations during brine

regeneration as high as 9.1 mg/L (see Table 3-13) (Sorg

et al., 1991). When the backwash water, brine waste, and

rinse water are combined, the average uranium concentration was in the range of 2 to 10 x 105 μg/L.

3.1.3.5 Iron Removal Processes Wastes

Oxidation/filtration and greensand filtration are basic

water treatment methods for iron and manganese removal, but field data have shown that these processes

Table 3-11. Summary of Radium Concentration in Lime Softening Sludges and Backwash Water 

Location/Waste

Source Watera

(pCi/L) Ra-226 Ra-228 Ra-226 pCi/L

Ra-228

pCi/L

Ra-226

pCi/g (dry)

Ra-228

pCi/g (dry)

W. Des Moines, IA

Sludge (clarifier drawoff)

Backwash water

Lagoon sludge

9.3

76.0

6.3b

5,159

—

—

596

—

—

10.8b

—

—

1.3

Webster City, IA

Sludge

Backwash water

6.1

980–1,114

50–92

—

—

—

—

—

—

Elgin, IL

Sludge (clarifier)

Sludge (clarifier)

Sludge (blanket)

Sludge (lagoon-active)

Sludge (lagoon-inactive)

Sludge (lagoon-entrance)

Backwash water

Sludge (filtrate)

3.5–7.5

—

948

—

 9,642

11,686

—

11.5–21.9c

0.5–0.48

—

 873

—

 9,939

12,167

—

—

—

2.8–10.7

8.6

1.3–12.5

11.3

10.9

6–30

—

—

—

8.0

—

11.7

11.3

—

—

—

Bushnell, IL

Sludge (clarifier)

12.6

4,577c — 21.6c —

Peru, IL

Backwash water

Sludge (pit)

3.1–6.1 —

9.6, 13.8, 87.7b

—

—

—

—

—

—

 9.2

—

—

—

Colchester, IL

Sludge (clarifier)

Backwash water

12.1

2,038

<20

236

<39

15.0

—

1.7

—

Beaver Dam, WI

Sludge (clarifier)

2.7–7.1 1.0–2.4

— — 1.4 1.4

Wapum, WI

Sludge (clarifier)

3.3–4.1 1.9–1.4

—

—

—

1.9

1.9

0.9

0.9

a Cornwell et al., 1987; AWWARF, 1969; U.S. EPA, 1987; U.S. EPA, 1976a; Myers et al., 1985; Sorg and Logsdon, 1980.

b Composite samples.

c

 Grab samples.

27
frequently remove some fraction of radium in the source

water as well (Bennett, 1978; Brink et al., 1978; U.S.

EPA, 1976a,b). These data show a wide range of radium

removal efficiencies, from almost 0 percent to as high

as 70 percent. These filtration processes produce a

backwash water waste stream from the filters; the concentration of radium in this stream will vary widely because of the wide range of removal. Table 3-14 shows

data from several filtration systems in Iowa, Illinois, and

Wisconsin. The concentration of radium in grab samples

of backwash water from these plants ranges from 120

to 1,980 pCi/L, while composite samples range from 34

to 165 pCi/L.

3.1.3.6 Reverse Osmosis and Electrodialysis

Wastes

RO and ED are capable of removing anions and cations

and, therefore, are effective for the removal of both

radium and uranium from backwash waste (Clifford,

1990; U.S. EPA, 1976a; Sorg and Logsdon, 1980; Sorg

et al., 1991). The quantity and concentrations of radioactive contaminants in reject water depend on the concentration of contaminants in the source water, the

Table 3-12. Summary of Radium-226 Concentrations in Brine Waste From Ion Exchange Treatment (Clifford, 1990; Bennett, 1978;

Brink et al., 1978; Clifford et al., 1988)

Location

Raw Water

Concentration 

(pCi/L)

Backwash Water

(pCi/L)

Brine

(pCi/L)

Avg. of Backwash

+ Brine + Rinse

(pCi/L)

Peak

Concentration in

Wastewater (pCi/L)

Eldon, IA 49 9–30a 400–3,500a — 3,500

Estherville, IA 5.7 94b 5–320a — 320

Grinnell, IA 6.7 6–19a

12b 210–320a — 320

Holstein, IA 13 7.8a 70–1,100a — 1,100

Dwight Correctional

Center, IL (Greensand)

 3.3 — — 22, 27, 29b —

Hersher, IL 14.3 — — 65–94b 315

Lynwood, IL 14.7 — — 64b

, 70b, 82b —

Elkhorn, WI 7.2 5.2–7.8a 27–158 23.1 158

Redhill Forest, CO 11–35 10–30 11–2,000 — 2,000

a Grab samples.

b Composite samples.

Table 3-13. Summary of Uranium Concentrations in Ion Exchange Treatment Plant Wastewater (Jelinek and Sorg, 1988; Sorg et

al., 1991)

Location

Raw Water

Concentration

(mg/L)

Backwasha

(mg/L)

Brinea

(mg/L)

Rinsea

(mg/L)

Avg. ofb

Backwash +

Brine + Rinse

(mg/L x 105

)

Peak

Concentration

in Wastewater

(g/L)

Arrowbear County

Water Distribution, CA

135 2.4–1,040 53.2–

6,780,000

36–

9,140,000

2.0–9.9 9.1

Coal Creek, CO 39–110 — — — 16.5 —

a Grab samples during regeneration cycle.

b Samples collected from storage tank.

Table 3-14. Summary of Radium-226 Concentrations in

Waste Stream From Iron Removal Filters

(Clifford, 1990; Bennett, 1978; Brink et al., 1978;

Clifford et al., 1988; U.S. EPA, 1992b; Wisconsin,

1984)

Location

Raw Water

Ra-226

(pCi/L)

Backwash

Water Ra-226

(pCi/L)

Eldon, IA (anthracite filter) 49 254–1,027a

Esterville, IA (anthracite filter) 5.7 165–1,980a

165b

Holstein, IA (anthracite filter) 13 80b

Adair, IA (greensand filter) 6.6 190a

Stuart (anthracite filter) 16 120–230a

Redhills Forest, CO

(conventional treatment package

plant)

31.5 60 pCi/L (filter

backwash)

Kaukauna, WI (greensand filter) — 52.5b

(anthracite filter) 33.5b

Elkhorn, WI (sand filter) 7.2 11.9 (avg.)

Herscher, IL (150 avg.)

(anthracite filter)

14.3 144b

, 159b, 149b

a Grab samples.

b Composite sample.

28
removal rate, and the fraction of water rejected. Assuming a high contaminant removal rate of 95 percent or

greater, a good approximation of the concentration of

radionuclides in reject water can be easily calculated by

the following:

Concentration in

reject water =

Concentration of radionuclide

in raw water

Fraction of water rejected

(Eq. 3-7)

Table 3-15 presents radium-226 concentrations in reject

waters from several RO plants in Iowa and Florida. The

data show a radium-226 concentration range of 7.8 to

43 pCi/L in the reject water.

Uranium levels in reject waters from a pilot plant study

conducted by Charlotte Harbour Water Association are

shown in Table 3-16. Reject water concentrations range

from 301 to 1,125 μg/L for three different RO membranes. This broad range is a result of the wide-ranging

feedwater uranium concentrations, 154 to 682 μg/L, and

the recovery rates of the individual membranes tested.

Although no field data on the concentration of radium or

uranium were found in the literature for ED, the concentration of these contaminants in the reject waste stream

from ED systems should be in the same range as RO

wastes because of the similarity in the treatment processes used for each method.

3.1.3.7 Waste Materials

Solid waste materials from drinking water treatment

processes can be grouped into two categories: 1) selective/specific sorbants that are not regenerated or

reused, and 2) materials used in treatment processes

that will retain some level of contamination and are

replaced periodically (e.g., filter sand). These materials

are listed in Table 3-10. Of the two categories, selective/specific sorbants will contain the highest concentration of contaminant by a very large factor because, by

design, they are used to remove specific contaminants

from source water. Two of the more widely known sorbants are the radium selective complexes (RSC), used

for radium removal, and GAC, used for radon removal.

The amount of radioactive contaminant retained on

Table 3-15. Summary of Radium-226 Concentrations in Reject Water of Reverse Osmosis Treatment

Location

System Capacity

(1,000 gpd)

Raw Water TDS

(mg/L)

Raw Water

Ra-226 (pCi/L)

Percent Water

Rejected (%)

Reject Water

Ra-226

(pCi/L)

Bay Lakes Estates MHP, FLa 40 2,532 3.2 — —

Venice, FLa 1,000 2,412 3.4 46 7.8

Sorrento Shores, FLa 200 3,373 4.6 61 7.9

Spanish Lakes MHP, FLa 70 1,327 10.4 69 20.5

Nokomis School, FLa 0.8 1,442 11.1 — 11.9

Bayfront TP, FLa 1.6 895 12.1 72 19.4

Kings Gate TP, FLa 30 1,620 15.7 — —

Sarasota Bay, FLa 5 2,430 20.5 50 37.9

Greenfield, IAb 150 2,200 7.6 31 43

Lamont, ILc,d

• Pilot system-1 — 510 11.6 50 19.2

• Pilot system-2 — 681 13.9 85 14.2

• Pilot system-3 — 451 13.0 75 14.1

a Sorg et al., 1980.

b U.S. EPA, 1976b.

c

 Clifford et al., 1988.

d Single membranes.

Table 3-16. Summary of Uranium Concentration in Reject

Water of Reverse Osmosis Treatment (U.S. EPA,

1988c)

Location

System

Capacity

(1,000

gpd)

System

TDS

(mg/L)

Raw

Water

Uranium

(mg/L)

Percent

Water

Rejected 

(mg/L)

Rejecta

Water

Uranium

Charlotte Harbour, FLb

Membrane 1 5 450 682 40 1,125

Membrane 2 5 760 154 50 301

Membrane 3 5 575 277 90 304

a Calculated based on 99 percent rejection.

b Single membrane.

Key

TDS = total dissolved solids.

29
these materials depends on the concentration of the

contaminant in the source water, the percent removal of

contaminants, and the length of operation (usually determined by the capacity of the sorbant).

RSC material has been used to remove radium directly

from source water and from the brine waste of IX processes (Clifford et al., 1988; U.S. EPA, 1988b; U.S. EPA,

1985a). Samples of RSC material from several WTPs in

Illinois, Wisconsin, and New Hampshire, show radium

concentrations ranging from 3.6 to 9.1 pCi/g (Redhill

Forest, CO) (see Table 3-17). The low value of 3.6 nCi/g

was the average of three RSC samples from an exhausted bed treating a water source with a total radium

concentration of 18.1 pCi/L. The higher concentration of

9.1 nCi/g was the calculated concentration of radium226 on an RSC system at 1 year, removing radium

(1,200 pCi/L) from a spent brine solution of an IX plant.

This system was not run to exhaustion at that point in time.

GAC is used to remove radon from source water, but

because of the short half-life of radon, the longer-lived

decay product, lead-210, is actually contained on the

GAC (U.S. EPA, 1990b; Lowry and Brandow, 1985;

Watson and Crawford-Brown, 1991). The amount of

lead-210 retained on GAC can be calculated from data

on the source water radon concentration and the percent removal.

Core GAC samples from two GAC radon removal plants

in New Hampshire were analyzed for lead-210 after 293

days of operation (Mont Vernon), and after approximately 4 months (Amherst). As expected, the concentration of lead-210 decreased with bed depth. The

maximum level at the top for the Mont Vernon system

was 757 pCi/g, and 297 pCi/g for Amherst. The Mont

Vernon system had a higher influent radon concentration (210,500 pCi/L) as compared with the Amherst system (41,800 pCi/L), and had been operating for

approximately 3 months longer. Thus, a higher level of

lead-210 would be anticipated at the Mont Vernon site.

Neither system was considered to be close to radon

exhaustion, each having an expected bed life of 5 to 10 years.

3.2 Physical Characteristics of Residuals

The increasing need to process WTP residuals in preparation for their final disposal/beneficial reuse has intensified investigations into their physical characteristics.

These characteristics significantly affect the ability to

handle, thicken, dewater, and convey WTP residuals

prior to the disposal/beneficial reuse. The following sections discuss solids content, specific resistance, comTable 3-17. Concentration of Radionuclides on Water Treatment Process Media and Materials (Bennett, 1978; Brink et al., 1978)

Location Treatment Process

Process

Media/Material

Radionuclide

Ra-226 Concentration

Lamont, IL (Pilot Study) Radium selective complexer Resin Ra-226 

Ra-226

3.6 nCi/g (exhaustion)

(total radium)

Redhill Forest, CO Radium selective complexer

(brine treatment)

Resin Ra-226 770x188 pCi/ft3

Herscher, IL Iron removal

Cation exchange

Filter media

Resin

Ra-226

Ra-228

Ra-226

111.6 pCi/g

38.9 pCi/g

43 pCi/g

Lynwood, IL Cation exchange Resin Ra-226

Ra-228

9.6 pCi/g

6.6 pCi/g

Dwight Correctional Institute, IL Natural greensand

(cation exchange)

Greensand Ra-226 29–46 pCi/g

Peru, IL Lime softening Filter media Ra-226

Ra-228

4.6 pCi/g

3.6 pCi/g

Elgin, IL Lime softening Filter media Ra-226

Ra-228

16.0 pCi/g

8.3 pCi/g

Elkhorn, WI Iron removal

Cation exchange

Filter media (sand)

Resin

Ra-226

Ra-228

Ra-226

Ra-228

1.47 pCi/g

0.48 pCi/g

6.04 pCi/g

2.7 pCi/g

Mt. Vernon, NH GAC 

Radon removal

GAC U-235/238

Pb 210

549–9,050 pCi/g (top

of bed)

757 pCi/g (max)

Amherst, NH GAC 

Radon removal

GAC Pb-210 297 pCi/g (max)

—

Key

GAC = granular activated carbon.

30
pressibility, shear stress, density, and particle size—

physical characteristics that can affect the solids dewatering process.

3.2.1 Solids Content

The solids content of residuals varies widely, depending

on factors such as the raw water characteristics, coagulant type and dose, and whether lime is used. The data

shown in Table 3-18 indicate the range of solids found

for a variety of water treatment sludges in Missouri. The

alum sludges have the lowest solids content, while the

lime sludges and sludges from Missouri River plants

have the highest. In general, the greater the amount of

incorporated suspended solids and calcium carbonate

in the sludge, the higher the solids content. Therefore,

when low-turbidity water is treated, the solids content of

the residuals will be low, while the solids content of

residuals from turbid waters will be much higher.

In Table 3-19, the data show that the solids content

following gravity settling and vacuum dewatering depends on the raw water turbidity, coagulant dose, and

coagulation mechanisms. Because the pH at which

coagulation occurs will determine the coagulation

mechanism, pH will also influence the residuals solids

content.

For example, in Table 3-19, the data show that the

thickened and dewatered solids content are lower when

coagulation is conducted at pH 8.1 than when it occurs

at pH 6.5. At a coagulation pH of 6.5, the predominant

means of incorporating the natural turbidity-causing particles into the sludge was adsorption-charge neutralization (i.e., the positive hydrolysis species formed by

adding alum to the water were able to adsorb on the

surface of the natural turbidity particle, neutralizing the

negative charge of these particles and allowing them to

be incorporated into the sludge). This process results in

a lower alum dose and, consequently, in a sludge that

is easier to thicken/dewater due to a higher percentage

of natural turbidity-causing particles and a lower percentage of gelatinous aluminum hydroxide.

At a pH of 8.1, the predominant means of incorporating

the natural turbidity into the sludge was enmeshment in

a precipitate, or sweep-floc (i.e., the alum added reacted

with the alkalinity of the water to form large amounts of

gelatinous aluminum hydroxide precipitate, in which the

natural turbidity particles were entrapped). This type of

sludge is more difficult to thicken/dewater because of

the high percentage of aluminum hydroxide present.

Also, alum sludge has been shown to have a lower

solids content than iron or lime sludge.

Table 3-18. Settled Solids Concentration of Residuals From

Water Treatment Plants in Missouri (Calkins and

Novak, 1973)

Location Sludge

Settled

Solids (%)

Higginsville Alum 3.1 

Macon Alum 3.4 

Boonvillea Lime and alum backwash 3.96

Jefferson Citya Lime and iron backwash 4.1 

St. Louis Countya Iron backwash 4.62

Moberly Alum 6.3 

Kirksville Alum 7.8 

Kansas Citya High magnesium softening sludge 8.0 

Boonvillea Lime and alum 8.17

St. Louisa Iron backwash 8.95

Boonville Lime and alum 10.1 

St. Josepha Catfloc backwash 11.3 

St. Louis Countya Iron (secondary basin) 12.2 

Kansas Citya Lime and magnesium 15.2 

Jefferson Citya Lime and iron 19.1 

St. Louisa Iron (secondary basin) 19.3 

St. Louis Countya Iron (primary basin) 21.1 

Kansas Citya Softening sludge 25.3 

Jefferson Citya Lime and iron 26.8 

Columbia Lime 33.0 

St. Louisa Lime and iron (primary basin) 35.6 

St. Josepha Catfloc 35.8 

Mexico Lime 54.0 

Table 3-19. Effect of Coagulation Mechanism on Alum Sludge Properties (Knocke et al., 1987)

Coagulation Conditions

Influent Turbidity

NTU

Coagulant

Dose mg/L

Coagulation

pH Mechanism

Ultimate Thickened

Solids Conc.

Percent

Specific

Resistance r

m/kg x 1011

Vacuum

Dewatered Solids

Conc. Percent

40 10 6.2 Adsorption-charge neutralization 6.0 55 42

40 15 6.3 Adsorption-charge neutralization 5.5 60 22

 7 40 6.5 Mixed 1.0 95 15

 7 75 7.1 Enmeshment 1.0 150 11

 7 75 8.1 Enmeshment 0.5 310 9

31
The solids content of residuals varies through a plant

and depends on the chemicals added and the addition

sequence. In Figure 3-2, the change in solids throughout

a Missouri River surface water treatment plant is shown.

The settling sludge, consisting of settled clay and calcium carbonate residues, settles to greater than 60 percent solids. At the other end of the plant, the settled filter

backwash is 4.6 percent solids. Filter backwash solids

generally will consist of the lighter solids that fail to settle

in sedimentation basins, so they are usually lower in

solids than is material from the clarifiers. As shown in

Figure 3-2, the decrease in settled solids corresponds

to a decrease in calcium carbonate as reflected by the

calcium-to-iron weight ratio of the sludge solids.

With regard to softening residues, calcium carbonate

solids form dense slurries with a high solids content.

Settled solids in excess of 60 percent have been found

for relatively pure calcium carbonate. When Mg(OH)2 is

incorporated into softening residues, however, the solids

content decreases, usually in direct proportion to the

magnesium fraction. Figure 3-3 shows that the settled

solids content of softening residues can vary from less

than 10 percent to more than 60 percent as the magnesium fraction in the sludge decreases.

An operational definition for water in sludges divides the

water into three simple fractions—nonremovable water,

capillary water, and removable floc water. While there is

considerable debate about the types and distribution of

water in sludges, this operational definition is useful in

estimating the potential solids content that can be

achieved by dewatering. Nonremovable water is the

water associated with surfaces and has been described

as “vicinal” water. This water is thought to form a hydration shell around surfaces. The large crystalline surfaces

of calcium carbonate residues will have relatively little

nonremovable or visual water because the surface area

of this material is relatively small.

Capillary water is that which is trapped between particles and retained by capillary action. Capillary water is

removed by application of pressure, which results in

flocs moving closer together and squeezing out water

between flocs. Removable floc water is water retained

within flocs in a manner similar to that of a sponge; it is

removed as the flocs are squeezed or compressed by

the application of pressure. In effect, both capillary water

and removable floc water are expelled simultaneously

by the application of pressure. As the pressure increases, more water is removed; at the highest pressure, any remaining water is primarily nonremovable

water. This process is depicted in Figure 3-4.

Applied pressure is the major operational variable in

dewatering. As the applied pressure of a dewatering

process is increased, the cake solids concentration

increases, but the lower-solids alum sludge is always

well below lime sludge levels, no matter what process

Figure 3-3. Effect of Ca-to-Mg ratio on the solids concentration

of softening sludge (Calkins and Novak, 1973).

Figure 3-2. Change in sludge settled solids concentration throughout a treatment plant (Calkins and Novak, 1973).

32
is used (see Table 3-20). Lime sludge particles contain

almost no floc water. Water retained after settling is fixed

to surfaces and between particles (capillary water).

Alum sludge, in addition to a having higher nonremovable water fraction, contains a large amount of internal

floc water.

Because polymer conditioning agents interact only with

the outer surface of flocs and do not alter their internal

floc structure, polymer conditioning neither affects internal floc water or vicinal water, nor significantly affects

dewatered cake solids. Therefore, the solids content of

any water treatment residue is determined by the inherent composition of the solids as they form and by the

applied pressure of the dewatering device. Once

formed, only processes that affect the internal floc structure (such as freeze-thaw conditioning) will substantially

influence the solids content that can be achieved by

dewatering.

3.2.2 Specific Resistance

Specific resistance is a measure of the rate at which a

sludge can be dewatered, and it reflects the size of

particles in the filter cake as water passes through the

sludge. While initially developed to assess vacuum filter

performance, specific resistance also indicates the dewatering rate by a variety of filtration processes, including gravity settling and sand bed dewatering.

A residue’s specific resistance to filtration is usually

measured using a vacuum filter device. The test is described in detail in Cornwell et al. (1987). Resistance to

filtration depends on the porosity or permeability of the

sludge cake. Permeability is a function of particle size

and particle deformation (compressibility) when pressure is applied. Specific resistance may be calculated

from filtration data using the following formula:

r = 2 PA2

b

μc

(Eq. 3-8)

where

r = specific resistance to filtration

P = pressure drop across sludge cake

A = surface area of filter

μ = filtrate viscosity

c = weight of dry solids deposited per volume of

filtration

b = slope of a plot of t/V versus V

t = time of filtrate

V = filtrate volume

Specific resistance data are useful for comparing

sludges and for evaluating the effect of polymers on

dewatering (See Figure 3-5). Because the specific resistance of a sludge often depends on the pressure

applied during filtration (see Section 3.2.3) and on the

mixing applied during chemical addition, resistance values are used primarily as an index of dewaterability

rather than as well-defined sludge properties. Sludge

Table 3-20. Specific Gravity of Sludge Particles and Cake Solids Concentrations Obtainable From Various Laboratory Dewatering

Methods (Novak, 1989)

Cake Solids Concentration (%)

Type of Slurry

Specific Gravity

of Particles

Settled Solids

Concentration (%)

Vacuum

Filtration Centrifugation

Pressure

Filtration

Lime sludge (low Mg) 1.19 28.5 56.1 60.6 69.5

Iron sludge 1.16 26.0 50.1 55.6 64.6

Ferric hydroxide 1.07 7.2 22.7 28.2 36.2

Lime sludge (high Mg) 1.05 5.6 21.0 24.8 34.6

Aluminum hydroxide 1.03 3.6 17.2 19.0 23.2

Figure 3-4. Water distribution and removal in a softening

(CaCO3) slurry and coagulant (Al(OH)3) slurry

(adapted from Novak, 1986).

33
resistances change during storage as a result of chemical aging and biological activity. Specific resistance values are most useful in comparing treatment options,

conditioning chemicals, or mixing conditions.

Specific resistance of sludges varies widely, as shown

in Table 3-21. In general, residuals with specific resistance of 10 x 10 m/kg or less are considered to dewater

readily, while those with a specific resistance of 100 x

1011 m/kg have poor dewaterability. As indicated by the

data in Table 3-22, lime softening sludges dewater rapidly and alum sludges from relatively clean surface waters dewater slowly.

Specific resistance for coagulant sludges increases as

the pH rises and as the raw water turbidity decreases.

The factors in coagulation that lead to high moisture

content in the sludge cakes also cause sludge to dewater slowly. Alum sludges from low turbidity raw waters

have both a low rate of dewatering and a low solids

content.

Chemical conditioners, usually anionic polymers, may

be added to sludges to decrease the specific resistance.

The selection of polymers and the appropriate dose for

conditioning is usually determined by laboratory testing.

Factors influencing the polymer dose include the solids

content, pH, and mixing intensity.

3.2.3 Compressibility

Most coagulant residues are highly compressible. Compressibility is thought to occur as a result of floc deformation during dewatering. The major consequence of

compression is that dewatering rates decrease as the

pressure applied during the process increases.

The compressibility of residues is often represented by

the following equation:

r2 = r1∆PS (Eq. 3-9)

where

r1 = specific resistance at pressure, P1

r2 = specific resistance at pressure, P2

∆P=P2 - P1

S = coefficient of compressibility

The coefficient of compressibility, S, is usually determined from a log-log plot of specific resistance (R)

versus pressure (P). For coagulant residues, values of

S vary from approximately 0.8 to 1.5, while values for

softening residues may be as low as 0.4 for materials

comprised primarily of calcium carbonate.

Similar relationships have been described, where compressibility relates to empirical constants that are determined experimentally. A somewhat different approach

describes the formation of a skin at the surface of the

filtering medium that resists the transport of water

through a filter cake. These approaches all recognize

Table 3-21. Specific Resistance for Various Chemical

Sludges (Calkins and Novak, 1973)

Location Sludge

Specific

Resistance 1010

(m/Kg)

Jefferson City Lime and iron 2.11

Jefferson City Lime and iron 4.3

Kansas City High magnesium softening

sludge

 5.49

Boonville Lime and alum 5.83

Boonville Excess lime and alum

backwash

 5.98

Jefferson City Lime and iron 6.12

Jefferson City Lime and iron 6.79

Jefferson City Lime and iron 7.0

Kansas City Softening 11.57

Boonville Excess lime and alum

backwash

 13.2

St. Joseph Cationic flocculent 14.1

St. Louis Lime and iron 21.2

Kansas City High magnesium softening

sludge

 25.1

St. Louis Iron 40.8

Boonville Lime and alum 53.4

St. Louis County Iron backwash 76.8

St. Louis County Iron 77.6

St. Joseph Cationic-flocculent

backwash

 80.1

St. Louis Iron backwash 121.8

St. Louis County Iron 148.5

Moberly Alum 164.3

Figure 3-5. Use of specific resistance to determine optimum

chemical dosage (Cornwell et al., 1987).

34
that residuals tend to deform under pressure, resulting

in an increased resistance to filtration.

Some of the changes in filtration resistance are attributable to particle shear, which results from the movement

of water through the cake. The particles in sludge cakes

undergoing filtration have been shown to desegregate,

an effect that was especially troublesome for alum

sludges. In contrast, calcium carbonate slurries were

resistant to shear. Because the breakup or disaggregation of flocs led to increased polymer conditioning requirements, high pressure dewatering systems may be

appropriate for alum sludges.

Generally, solids that consist primarily of coagulant materials from clean raw water sources dewater poorly

when pressure is applied. In contrast, solids with a large

fraction of rigid particles, such as softening sludges, will

not deteriorate as dramatically when pressure is applied. Whether this effect is caused by particle deformation, skin formation, or particle breakup may be less

important than the effect high compressibility has on

dewatering and on the selection of dewatering equipment.

3.2.4 Shear Stress

Shear stress is an important characteristic in determining the handleability of a sludge. The undrained shear

strength of various water treatment residues, shown in

Figure 3-6, varies markedly with the solids content.

Figure 3-6 also shows that the sludge settled solids

concentration provides a reasonable estimate of the

range of solids concentrations where a sludge makes

the transition from a liquid to a handleable solid. This

condition is clearly presented in Figure 3-7, where the

solids concentration needed to produce a handleable

sludge occurs in a range of 0.02 to 0.05 tons/ft2

. The

data in Figures 3-6 and 3-7 also show that alum sludges

Table 3-22. Summary of Floc Density and Dewatered Solids Concentration Data for Several Chemical Sludges (Knocke et al., 1993)

Maximum Dewatered Solids Concentration (% by wt)

Sludge Type

Floc Density

(g/mL)

Dry Solids

Density (g/mL) Grav. Thick. Vacuum Filter Centrifuge Pres. Filt.

Polymer WTP 1.06–1.08 1.58 2.7 17 11 21

Polymer WTPa >1.30 1.60 25 45 33 40

Iron I 1.26–1.28 2.86 8.9 27 31 30

Iron I >1.30 2.86 16 54 44 53

Alum I 1.22–1.25 2.55 3.0 17 15 25

Alum III 1.14–1.16 2.45 3.0 14 10 —

Lime >1.30 2.47 13 41 42 —

a Freeze-thaw conditioning provided.

Figure 3-6. Variation in shear strength with sludge moisture content (Novak and Calkins, 1973).

35
generally fall in the settled solids range of 7 percent and

below. Therefore, solids concentrations of 15 to 20 percent may be sufficient to produce a handleable sludge.

In contrast, some softening sludges may require concentrations above 50 percent before they can be handled.

In a study of the dry weight density-moisture relationship

for an iron coagulant and two alum coagulant sludges,

the iron sludge showed the typical humped curve, where

maximum density occurs at an optimum moisture content. The alum sludge reached a maximum density at

the lowest moisture content (Figure 3-8). The coagulant

residuals were extremely plastic and compressible and

greatly exceeded these values for high clay soils.

When landfilling these materials, the solids content

should be as high as possible to minimize the amount

of landfill space required. In addition, because the cost

of landfilling is most often determined on a weight basis,

increasing the solids content and reducing the water

content of residuals prevents having to pay for the cost

of landfilling water.

3.2.5 Density

Floc density varies with floc size, with density decreasing as floc size increases (see Figure 3-9). The major

impact of mixing shear is to make flocs smaller. For

similar floc sizes, mixing has no effect on floc density, a

finding supported by others.

As the volume of suspended solids (Kaolinite clay) in the

floc increases, the floc density also increases. Settled

and dewatered cake solids increase as the suspended

solids in the cake increase. The effect of the solids-tocoagulant ratio on floc density and cake solids suggests

that these two factors are related. The apparent density

or specific gravity of flocs is a useful predictor of the

dewatered cake solids produced by a variety of processes (Table 3-20), with floc density measurements

varying from 1.03 to 1.19 for various water treatment

residuals. In a study of the effects of organic matter on

floc density, when flocs contained more organic matter,

0 20 40 60

0 100 200 300

0 80 160 240

Water Content (%)

SLUDGE 3


ALUM

SLUDGE 2


ALUM

SLUDGE 1


IRON

Dry Unit Weight (KN/m3) 16


14


12


10


8

12


8


4


0

11


9


7


5


3

Figure 3-8. Compaction curves of test sludges (Cornwell et al.,

1992).

Figure 3-7. Comparison of sludge settled solids concentration

with the solids concentration where a sludge becomes “handleable” (Novak and Calkins, 1973).

36
their density declined, dewatering rates decreased and

dewatered cake solids decreased (see Figures 3-10,

3-11, and 3-12). Floc densities for alum sludges have

been reported in the range of 1.14 to 1.22, depending

on the amount of total organic content (TOC) incorporated in the flocs.

A recent study comparing various methods of floc density measurement showed that the use of a low osmotic

pressure gradient medium to measure floc density is

more sound than other commonly used methods such

as the sucrose method. High osmotic pressure media

are likely to produce densities higher than actual because the high osmotic pressures cause the flow of

water out of the floc during measurement. The data

presented in Table 3-22 show floc densities ranging from

1.06 to greater than 1.3 g/mL for various water treatment

residues. The solids content obtained by various dewatering methods suggests that floc density may be an

important determinant in the cake solids obtained by

dewatering.

3.2.6 Particle Size Distribution

According to the theory of filtration, the resistance of

sludges to filtration is a function of particle size of the

flocs in the sludge cake. While several factors such as

compressibility and filter media blinding may cause variation from a precise relationship between particle size

Figure 3-11. Effect of incorporation of organic carbon on the

relative size distribution of aluminum hydroxide

sludge floc formed at pH 6.5 (Dulin and Knocke,

1989).

Figure 3-12. Variations in measured floc density as a function

of both coagulation pH and presence or absence

of TOC from sludge floc matrix (Dulin and Knocke,

1989).

Figure 3-10. Variations in dewatered cake solids concentration

of aluminum hydroxide sludges as a function of

organic content (Dulin and Knocke, 1989).

Figure 3-9. Variation of floc density with floc size (Lagvankar

and Gemmell, 1968).

37
and cake resistance, measurements of particle size

generally support this theory.

Data presented in Figure 3-13 for various metal hydroxide sludges show the relationship between mean floc

size and specific resistance. Additional data shown in

Figure 3-14 for alum sludges show a similar trend.

These data were measured using a HIAC particle

counter and indicate that unconditioned alum sludge

has a mean floc size of 20 μ or less.

Conditioning chemicals can also be seen to influence

particle size. Data presented in Figure 3-15 show the

increase in particle size resulting from addition of polymer and the associated decrease in the specific resistance to filtration.

Although the mean particle size is the primary factor in

determining sludge filtration behavior, two other factors

are important. If the particle distribution is bimodal, the

sludge is susceptible to “blinding,” which is defined as

the migration of fines through the cake, resulting in much

lower cake permeability near the filtering surface. This

phenomenon has been documented for certain sewage

sludges but does not appear to be common for WTP

residuals.

A more likely problem, especially with alum sludges, is

the formation of small particles from the breakup or

disaggregation of alum floc due to shear. Alum sludge is

very sensitive to shear (Figure 3-16), and shear, G,

equal to 500/sec can be attained in filter cakes from the

passage of water during vacuum dewatering. Therefore,

much of the demand for conditioning chemicals results

from the shear associated with the dewatering process

and not because alum sludge is comprised of small

particles. An important role of conditioning chemicals is

to make the sludge resistant to shear.

Particle size, as measured by conventional commercial

particle counters, generally verifies that particle counting

is a useful means to evaluate sludge dewatering properties. Using microscopic examination and considering

the particles to be elliptical, the relationship between

Figure 3-13. Floc size and resistance of metal hydroxide sludges to dewatering by vacuum filtration (Knocke et al., 1980).

Figure 3-14. Effect of specific surface area on the specific resistance of alum sludges (Knocke et al., 1980).

38
particle surface area and specific resistance has been

described.

3.3 Chemical Characteristics of

Residuals

The chemical characteristics of WTP residuals tend to

affect the options for disposal/beneficial reuses more

than they affect the ability to handle, thicken, or dewater

residuals.

3.3.1 Solids Content

The solids content of WTP residuals varies widely,

based on whether the residual is a liquid waste, sludge

(i.e., semi-solid waste), or solid waste. Furthermore, the

solids content of sludges varies significantly depending

on the solids handling processes to which the sludge is

subjected (e.g., thickening, dewatering).

The type and concentration of solids affect the distribution of water within a sludge (Cornwell et al., 1987).

Because floc water is trapped within the flocs, capillary

water is held to sludge flocs by surface tension and

attractive forces, and bound water is chemically bound

to individual floc particles. Equally important to the solids

content is the residual’s volatile solids-to-total solids

ratio (i.e., VS/TS). Fortunately the majority of the solids

in WTP residuals tend to be inert, and the VS-to-TS ratio

is typically less than 30 percent. In coagulant sludges,

the inert SS tend to be aluminum or iron hydroxide

precipitates that can be difficult to dewater because of

their gelatinous nature. In softening sludges the inert SS

are associated with crystalline calcium carbonate, which

drastically reduces the flow water trapped within the floc.

3.3.2 Metals Content

The metals content of WTP residuals is important for a

number of reasons: 1) potential impacts on the disposal

of the residual in a sanitary landfill; 2) possible inhibitory

effects if the residuals are discharged to a WWTP for

processing; 3) potential adverse contributions to the

residuals from the WWTP based on Part 503 sewage

sludge regulations; and 4) possible effects on the whole

effluent toxicity of the effluent from the WWTP.

The mean total levels of cadmium, copper, chromium,

nickel, lead, and zinc in coagulant sludges from WTPs

are generally 10 to 35 percent of the corresponding

values for sewage sludges. Except for cadmium, 76 to

Figure 3-16. Effect of Gt on optimum polymer dose for alum

sludge conditioning (Werle et al., 1984).

Figure 3-15. Representative results from metal hydroxide sludge conditioning studies (Knocke et al., 1980).

39
87 percent of these were found to be within the oxide or

silicate matrix of the alum and iron sludges. Although

cadmium could become mobilized under acidic conditions, the levels were measured at levels too low to

promote significant leaching of the cadmium. Another

investigation into the mobilization of several heavy metals from ground-water and surface water sludges suggests extreme decreases in pH (i.e., less than 2.5) and

alternating aerobic/anaerobic conditions are necessary

for significant mobilization.

Fortunately, it is known that much of the heavy metals

content of WTP sludges is often contributed by impurities in the coagulant. Consequently, heavy metal concentrations in the residuals can be limited by carefully

specifying the coagulants and other chemicals added to

the water.

3.3.3 Toxicity

Before 1990, the potential toxicity of WTP residuals was

determined based on the EP toxicity test. The Toxicity

Characteristic Leaching Procedure (TCLP) has now replaced the EP toxicity test. Coagulant sludges from

WTPs have been shown to easily meet the TCLP criteria. The metals content of WTP residuals is not anticipated to be a problem under the criteria for the TCLP.

40
Chapter 4

Water Treatment Residuals Processing

Residuals handling at water treatment plants (WTPs)

has traditionally dealt with the handling of waste streams

from sedimentation, precipitation, and filtration from

conventional coagulation type plants or lime softening

facilities. Historically, residuals from these types of facilities have been most commonly disposed of through

discharge to sanitary sewers, streams, or similar bodies

of water. The changing regulatory environment is leading to an increase in the number of plants incorporating

solids handling facilities. This chapter provides information concerning the selection of residuals handling processes for a WTP, including:

• Basic descriptions of standard residuals handling

processes.

• Definition of preliminary residuals processing requirements.

• Criteria for the preliminary selection of unit process

combinations.

• Discussion of sizing procedures for select unit processes.

This chapter identifies unit processes for treating traditional types of residuals. Changes in the regulatory environment, however, are also requiring an additional

form of residuals handling—that of air emissions control.

This form of control may be applied to the discharge of

gaseous residual byproducts from processes such as

ozonation or air stripping.

4.1 Residuals Handling Process Types

Some water treatment processes that may produce solids are grit collection, sedimentation, and filtration (Figure 4-1). These processes can include mechanisms for

collection and concentration of solids before conveyance to disposal or to another unit process. These residuals can be handled through a variety of process

types. A residuals handling flow schematic with each of

the process types is shown in Figure 4-2.

The process types are:

• Thickening: A process of concentrating the solids

content of a residual stream to reduce the volume

before disposal or further treatment.

• Coagulant recovery: A treatment technique for improving solids dewatering characteristics and lowering the concentration of metallic ions in the residuals.

Recalcination is a related process associated with

lime softening sludges (see Chapter 13).

• Conditioning: Adding a chemical to a residual or

physically altering its nature. Conditioning is traditionally used as a method to optimize the dewatering

process.

• Dewatering: Similar to thickening in that both processes involve a liquid-solids separation approach

with a goal of minimizing the amount of residuals for

disposal. Dewatering is defined as a process to increase the solids concentration of residuals (by

weight) to greater than 8 percent, typically in the 10

to 20 percent range.

• Drying: An extension of the liquid-solids separation

approach of thickening and dewatering. Drying is defined as a process to increase the solids concentration of residuals (by weight) to greater than 35

percent.

• Disposal and reuse: Removal of residuals from the

WTP site or permanent storage of residuals at the

WTP site. This category includes hauling to landfill,

discharging to sanitary sewer or natural waterway,

land application, and various reuse options (e.g., soil

supplement, brick manufacture).

• Recovered and nonrecovered water handling: Thickening, dewatering, and drying processes produce

both liquid and solids components. The solids component may be further treated and disposed of. The

liquid component is returned to the main WTP processes if it is recoverable, which means it has little

impact on the main treatment process and no harmful

effect on the finished water quality. Quality parameters that can affect the recoverable status of the water

include the following:

– Residuals metal concentrations.

– Disinfectant byproduct formation potential.

– Use of unapproved polymers in the residuals handling processes.

41
Nonrecoverable water must either be disposed of or

subjected to further treatment.

• Other processes: Those that do not readily fit into a

category listed above, including equalization, chemical conditioning, and residuals conveyance.

Actual residuals handling facilities may use any or all of

these processes in different combinations. Figures 4-3A

and 4-3B depict a facility that uses collection, thickening,

and dewatering processes. An example of a facility with

only one process in operation would be a plant with a

sedimentation basin and no sludge collection system.

The basin is taken out of operation periodically to manually remove the solids.

4.2 Process Descriptions

4.2.1 Collection Processes

Collection processes are the means by which WTP

residuals are collected from the process unit in which

they were removed from the water. In the water treatment process, residuals are removed from the process

stream by several different mechanisms. Inlet screens

remove larger pieces of debris (greater than 1 inch) from

the raw water source. Grit basins collect the coarsest,

densest material from the raw water source prior to

presedimentation. Presedimentation basins collect the

denser solids that do not require coagulation and flocculation for solids separation. Sedimentation basins promote gravity settling of solids particles to the bottom of

Raw


Water

Grit

Pre-


Sedimentation 


Waste

Rapid


Mix

Flocculation

Final Sedimentation


Basin

Coagulant


Sludge

Filter


Backwash


Waste

Finished 


Water

Filters Pre-


Sedimentation 


Basin

Figure 4-1. Residuals sources in water treatment plants.

Figure 4-2. Residuals handling process categories.

42
a water column where accumulated solids are then

removed. Several different types of processes can be

used in the sedimentation mode, such as chain and

flight, suction, and circular collector units.

In the conventional treatment plant, filtration is generally

the last step in the removal of suspended solids. Solids

are removed by a bed of granular media (sand, anthracite, and/or garnet) via straining, impingement, gravitational settling, or adsorption. Solids are removed from

the bed through a backwash procedure.

4.2.2 Thickening

Residuals concentration, or thickening, processes begin

after clarification, sedimentation, filtration, or water softening processes. Concentration processes are critical

to the economical removal of solids from the treatment

process. Thickening has a direct effect on downstream

processes such as conditioning and dewatering, and

can make the difference between an efficient, economical operation and an inefficient, high-cost one. WTP

residuals are most commonly concentrated using gravGrit


Basins

Cyclone Separator

Effluent

Grit

Grit Classifier

Filters Equalization


Basin

Presedimentation


Basins

UWRS


Splitter Box

Gravity


Thickeners

Presedimentation


Splitter Box

To 


Presedimentation


Basin

Recovered


Water

To Dewatering Process

Final 


Sedimentation


Basins

Grit Flow

Filter Backwash Flow

Sedimentation Basin Used Water Flow

Figure 4-3A. Residuals handling process schematic: sedimentation basin used water flow (Peck et al., 1993).

Gravity


Thickeners

Thickened


Sludge Thickened


Sludge


Equalization 


Basins

Dewatering


Centrifuges

Interim


Dewatering


Centrifuges

Evaporation


Lagoons


(Non-recoverable


Water)

Centrate Centrate

(Recoverable Water)

Centrate

Partially Dewatered

Solids (If Necessary)

Solids

Solids

Solids

Decant

(Recoverable Water) (Non-recoverable Water)

Decant

Solar Drying


Beds/Extended


Drying Area

To Ultimate


Disposal


(Off Site)

Figure 4-3B. Residuals handling process schematic: solids dewatering (Peck et al., 1993).

43
ity thickeners, but they can also be concentrated in

flotation thickeners or by gravity belt thickeners.

4.2.2.1 Gravity Thickening

Thickening of WTP residuals is most commonly performed in gravity thickeners, which work only when the

specific gravity of the solids is greater than 1. In this

process, both carbonate and metallic hydroxide residuals are conveyed to gravity settling tanks at a flow rate that

allows the residuals sufficient retention time to settle.

Gravity thickeners can be either batch feed or continuous flow. Residuals thickened in gravity thickeners may

require conditioning.

Thickener tanks (Figure 4-4) are generally circular and

are usually concrete, although small tanks can be made

from steel. They are typically equipped with rake mechanisms to remove solids. The floors are conical in shape

with a slope of between 10 and 20 percent. This slope

enables the mechanism to more efficiently move solids

to the discharge hopper.

Metallic hydroxide residuals, which come from either

clarifier operations or backwashing of filters, thicken to

only approximately 1 to 3 percent solids at loadings of

4.0 lb/day/ft2

. The degree of thickening is generally dependent on the hydroxide-to-total suspended solids

(TSS) ratio; high TSS solids can thicken to 5 to 30

percent solid. Carbonate residuals produced from water

softening processes settle readily and will thicken to

concentrations ranging from 15 to 30 percent solids at

loadings of 20 to 40 lb/day/ft2

 (Cornwell et al., 1987).

In a continuous feed thickening operation, the solids

slurry enters the thickener through ports in a column

located in the center of the tank. In theory, the solids are

distributed equally, both horizontally and vertically. The

solids settle to the bottom of the unit and the clarified

supernatant flows over discharge weirs located on the

periphery of the tank. These units are equipped with a

bottom scraper mechanism that rotates slowly, directing

the sludge to the drawoff pipe or sump near the bottom.

The thickener bottom is sloped to the center to help

collect the sludge. The slow rotation of the scraper also

prevents bridging of the solids.

Batch fill thickening tanks are often equipped with bottom hoppers. Sludge flows into these tanks, usually from

a batch removal of solids from the sedimentation basin,

until the thickening tank is full. The slurry is allowed to

settle in a telescoping decant pipe, which may be continuously used to remove supernatant. The decant pipe

may be lowered as the solids settle, until the desired

solids concentration is reached, or until the slurry will not

thicken further. The solids are then pumped out of the

bottom hoppers for further treatment or disposal.

4.2.2.2 Flotation Thickening

Flotation thickening is a solids handling option for residuals concentrates consisting of low-density particles.

Potential benefits include lower sensitivities to changes

in influent feed solids concentration and solids feed rate.

The process may also have more applicability for

sludges with high hydroxide components (greater than

40 percent by weight). While flotation has been used in

the European water industry, it has not been used for

long-term, large-scale thickening in the United States

(Cornwell and Koppers, 1990). The process is attracting

Figure 4-4. Gravity thickener cross-section (U.S. EPA, 1979b).

44
interest currently as both a concentration process and

as a thickening process in the water treatment industry.

Flotation thickening can be performed through any of

three techniques.

• Dissolved air flotation (DAF): Small air bubbles (50

to 100 μm in diameter) are generated in a basin as

the gas returns to the vapor phase in solution after

having been supersaturated in the solution.

• Dispersed air flotation: Large gas bubbles (500 to

1,000 μm in diameter) are dispersed in solution

through a mixer or through a porous media.

• Vacuum flotation: Operates on a similar principle to

the DAF with the condition of supersaturation being

generated through a vacuum.

Each of these techniques uses air bubbles to absorb

particles which may then be floated to the water surface

for separation from the liquid stream. DAF is the most

typically used flotation system in the municipal wastewater industry (U.S. EPA, 1979b).

In the DAF thickening process, air is added at pressures

in excess of atmospheric pressure either to the incoming

residual stream or to a separate liquid stream. When

pressure is reduced and turbulence is created, air in

excess of that required for saturation at atmospheric

pressure leaves the solution as the 50 to 100 μm-sized

bubbles. The bubbles adhere to the suspended particles

or become enmeshed in the solids matrix. Because the

average density of the solids-air aggregate is less than

that of water, the agglomerate floats to the surface.

Water drains from the float and affects solids concentration. Float is continuously removed by skimmers.

Usually a recycle flotation system is used for concentrating metallic hydroxide residuals. In this type of system,

a portion of the clarified liquor (subnate) or an alternate

source containing only minimal suspended matter is

pressurized. Once saturated with air, it is combined and

mixed with the unthickened residual stream before it is

released to the flotation chamber. This system minimizes high shear conditions, an extremely important

advantage when dealing with flocculent-type residuals

such as metallic hydroxides.

Several sources indicate that European facilities have

had success in concentrating hydroxide sludge to levels

between 3 to 4 percent solids (ASCE/AWWA, 1990;

Brown and Caldwell, 1990). These results appear to

include facilities that use flotation both as a concentration process (as an alternative to sedimentation) and as

a thickening process. Loading rates for hydroxide

sludges vary from 0.4 lb/ft2

/hr to 1.0 lb/ft2/hr for facilities

achieving from 2 to 4 percent float solids concentration.

Hydraulic loading of DAF units is reported at less than

2 gpm/ft2

 (Cornwell and Koppers, 1990).

The lack of historic operating data for this process indicates the need for bench- and pilot-scale testing prior to

selection of the process.

4.2.2.3 Gravity Belt Thickeners

Although the thickening of hydroxide slurries can be

accomplished using gravity belt thickeners, this technology is just beginning to be used. In this process, metallic

hydroxide sludge is discharged directly onto a horizontal, porous screen (Figure 4-5). As the sludge moves

along the length of the screen, water is removed by

gravity. Solids concentrations of 2.5 to 4.5 percent can

be achieved using these thickeners. The gravity belt

thickening units are made up of a sludge inlet port,

drainage screen, scraper blades, discharge outlet for

water, and a discharge outlet for the thickened residuals.

4.2.2.4 Other Mechanical Thickening Processes

Although gravity belt thickening is the most common

thickening process, some mechanical devices used for

dewatering may also be applied to this process. Examples include the continuous-feed polymer thickener,

drum thickener, and centrifuges. Solid-bowl-type centrifuges have been used in several pilot-scale studies

evaluating residuals thickening. Full-scale operating

data on mechanical processes for water treatment solids thickening are not currently available.

4.2.3 Conditioning

Conditioning is a process incorporated into many residuals handling systems to optimize the effectiveness of the

dewatering process. Conditioning of WTP residuals is

generally done by either chemical conditioning or physical conditioning.

4.2.3.1 Chemical Conditioning

Chemical conditioning is included in most mechanical

thickening or dewatering processes. This conditioning

involves the addition of ferric chloride, lime, or polymer.

The type and dosage of chemical conditioner vary

widely with raw water quality, chemical coagulants, pretreatment, desired solids concentration, and thickening/dewatering process used. A recent publication

(Malmrose and Wolfe, 1994) identified typical ranges of

conditioner use for hydroxide sludges in various mechanical dewatering systems (see Table 4-1). Conditioning agents used for lime sludges are typically lower in

volume if used at all. Recorded use of conditioning

agents for solids dewatered in open air processes is also

minimal.

A wide variety of polymers are available for use in the

dewatering processes. The most successful polymers

used are anionic with a high molecular weight. Polymers

can be obtained in a variety of dry and liquid emulsion

45
forms. A system should be equipped to feed either form

(Malmrose and Wolfe, 1994). Chemical suppliers should

be asked for a chemical’s NSF rating if process byproduct water is to be returned to the treatment plant

stream.

4.2.3.2 Physical Conditioning

The following physical processes may also be used to

optimize thickening/dewatering effectiveness (Cornwell

and Koppers, 1990):

• Precoat or nonreactive additives: Some dewatering

systems, primarily vacuum filtration and pressure filters, use a precoat additive in the process, typically

diatomaceous earth.

• Freeze-thaw conditioning: This process may be accomplished through either an open-air process in cold

weather climates or through mechanical equipment.

• Thermal conditioning at high temperatures (350°F to

400°F) under high pressure (250 to 400 psig): This

conditioning process is typically effective when there

is a high organic content present in the solids.

4.2.4 Dewatering

4.2.4.1 Air Drying Processes

Air drying refers to those methods of sludge dewatering

that remove moisture either by natural evaporation,

gravity, or induced drainage. Most air drying systems

were developed for dewatering residuals produced from

wastewater treatment plants (WWTPs) but have since

been used for the dewatering of WTP residuals. Air

drying processes are less complex, easier to operate,

and require less operational energy than mechanical

systems. They are used less often, however, because

they require a great deal of land area, are dependent on

climatic conditions, and are labor intensive. The effectiveness of air drying processes is directly related to

weather conditions, type of sludge, conditioning chemicals used, and materials used to construct the drying bed.

Sand Drying Beds

Sand beds are commonly used to dewater WTP residuals and have been used successfully for many years.

Dewatering on the sand bed occurs through gravity

drainage of free water (interstitial water in the residuals

slurry), followed by evaporation to the desired solids

concentration level. Figure 4-6 illustrates details of a

typical sand bed. In areas of high precipitation, covered

sand beds have been used.

Residuals on sand beds dewater primarily by drainage

and evaporation. Initially, water is drained through the

material, into the sand and removed through underdrains. This step, normally a few days in duration, lasts

Woven Synthetic Fiber Belt

Support Rollers

Pressure Belt

Pressure Belt Hydraulic Cylinder

Flexible 


Scraper


Blade

Safety


Shut Down


Monitor

Conveyor

Filtration and Washwater Discharge Upper Drain Pan Belt Wash Spray Nozzles

Air Actuated Pinch Rollers Bottom Drain Pan

Compressed Air

Flocculator

Rubber 


Covered


Drum

Figure 4-5. Gravity belt thickener cross section (Infilco Degremont, 1994).

Table 4-1. Typical Ranges of Conditioner Use for Hydroxide

Sludges in Various Mechanical Dewatering

Systems (after Malmrose and Wolfe, 1994)

Filter Press Centrifuge

Belt Filter

Press

Pyash precoat 10 lb/100 ft2 — —

Diatomaceous

earth precoat

6 lb/100 ft2 — —

Lime 10–30% — —

Ferric chloride 4–6 — 1–3

Polymer 3–6 2–4 2–8

Note: All values are in units of lb/ton dry solids unless noted otherwise.

46
until the sand is clogged with fine particles or until all the

free water has drained away. A decanting process removes any surface water. This decanting step can be

especially important for removing rain which, if allowed

to accumulate on the surface, can slow the drying process. Water remaining after initial drainage and decanting

is removed by evaporation.

Sand beds are more effective for dewatering lime residuals than residuals produced by coagulation with

alum. In both cases, however, conditioning the residual

slurry before discharging it into the sand bed helps the

dewatering process.

When designing a sand bed, the following factors should

be considered:

• Required solids concentration of the dewatered residuals.

• Solids concentration of the residual slurry applied to

the bed.

• Type of residuals supplied (lime or alum).

• Drainage and evaporation rates.

The required solids concentration depends on the technical or regulatory requirements for final residuals

disposal.

The amount of water that can be removed by drainage

is strongly influenced by the type of residuals applied to

the bed. The rate of evaporation varies with local climatic conditions and the solids surface characteristics.

Seasonal evaporation rates can be obtained from local

pan values. These values are tested and recorded by

the National Weather Service as measures of the local

evaporation rate. Because the crust that forms on the

surface of the sand bed inhibits evaporation, the pan

values must be adjusted when designing the sand bed.

An adjustment factor of 0.6 was experimentally derived.

Once cracking of the surface occurs, the evaporation

rate should again approach the pan value.

Thin layers of solids dry faster than a thick layer, but the

annual solids loading is of the depth of the individual

layers applied. Using too thin a layer has several disadvantages, including more frequent operation and maintenance, greater sand loss from the bed, and increased

costs.

To keep operation and maintenance costs as low as

possible, the design goal is to achieve the maximum

solids loading with the minimum number of application

and removal cycles.

Freeze-Assisted Sand Beds

Alum residuals have a gelatinous consistency that

makes them extremely difficult to dewater. By freezing

and then thawing alum residuals, the bound water is

released from the cells, changing the consistency to a

more granular type of material that is much easier to

dewater.

Freezing alum residuals changes the structure of the

residuals slurry and the characteristics of the solids

themselves. In effect, the solids matter is compressed

into large discrete conglomerates surrounded by frozen

water. When thawing commences, drainage occurs instantaneously through the large pores and channels

created by the frozen water. Cracks in the frozen mass

also act as conduits to carry off the melt water.

Freezing can be done mechanically or naturally. Because of the high cost associated with mechanical systems, natural systems are used most frequently.

The maximum potential response during both the freezing and thawing portions of the cycle can be obtained

by exposing the solids on uncovered beds. The drainage

water during thawing may move at a faster rate, and will

produce a greater volume than if applying the same

unconditioned solids to a conventional sand bed.

The critical operational requirement is that the solids

layer be completely frozen before the next layer is applied. Hand probing with a small pick or axe is the

easiest way to determine if this has been accomplished.

Solar Drying Beds

Until recently, paved beds used an asphalt or concrete

pavement on top of a porous gravel subbase. Unpaved

areas constructed as sand drains were placed around

the perimeter or along the center of the bed to collect

and convey drainage water. The main advantage of this

approach was the ability to use relatively heavy equipment for solids removal. Experience showed that the

Figure 4-6. Sand drying bed section (U.S. EPA, 1979b).

47
pavement inhibited drainage, so the total bed area had

to be greater than that of conventional sand beds to

achieve the same results in the same period.

Recent improvements to the paved bed process include

a tractor-mounted horizontal auger, or other device, to

regularly mix and aerate the sludge. The mixing and

aeration break up surface crust that inhibits evaporation,

allowing more rapid dewatering than conventional sand

beds.

Vacuum-Assisted Drying Beds

This dewatering technology applies a vacuum to the

underside of rigid, porous media plates on which

chemically conditioned residuals have been placed. The

vacuum draws free water through the plates, retaining

sludge solids on top and forming a cake of fairly uniform

thickness.

Cake solids concentrations of 11 to 17 percent can be

obtained on a vacuum-assisted drying bed, depending

on the type of solids being dewatered and the kind and

amount of conditioning agents used.

Problems with this method stem from two sources: improper conditioning and plate cleaning. The wrong types

of polymer, ineffective mixing of polymer and solids

slurry, and incorrect dosage result in poor performance

of the bed. Overdosing of polymer may lead to progressive plate clogging and the need for special cleaning

procedures to regain plate permeability. Plate cleaning

is critical. If not performed regularly and properly, the

media plates are certain to clog and the beds will not

perform as expected. Costly, time-consuming cleaning

measures are then required. Removal of dewatered

solids tends to be a constant, time-consuming operation.

Wedgewire Beds

The wedgewire, or wedgewater, process is physically

similar to the vacuum-assisted drying beds. The medium

consists of a septum with wedge-shaped slots approximately 0.01 inches (0.25 mm) wide. This septum serves

to support the sludge cake and allow drainage through

the slots (Figure 4-7). Through a controlled drainage

process, a small amount of hydrostatic suction is exerted on the bed, thus removing water from the sludge.

4.2.4.2 Lagoons

Lagoons are one of the oldest processes used to handle

water treatment residuals. Lagoons can be used for

storage, thickening, dewatering, or drying. In some instances, lagoons have also been used for final disposal

of residuals.

The lagoon process involves discharging residuals into

a large body of water. Solids settle to the bottom and are

retained in the lagoon for a long period. Sedimentation

and compression are two mechanisms used to separate

the solids from the liquid. Liquid can be decanted from

various points and levels in the lagoon. Evaporation may

also be used in the separation process if the residuals

are to be retained in the lagoon for an extended period.

The traditional lagoon consists either of earthen berms

built on the ground surface, or of a large basin excavated from the ground. Various types of systems are

installed in lagoons to decant the supernatant and, ultimately, drain the lagoon. State and local regulations

have become more stringent about preventing ground

pollution, and in some areas, laws affect the design of

WTP residual lagoons. Liners made of high-density

polyethylene (HDPE), leachate collection systems, and

monitoring wells are becoming common features of lagoon designs (see Figure 4-8). Lagoon depth typically

varies from 4 to 20 feet and the surface area ranges from

0.5 to 15 acres (Cornwell et al., 1987).

The effectiveness of lagoons in concentrating solids

typically depends on the method of operation. For metal

hydroxide solids retained in a lagoon for 1 to 3 months,

operating the lagoon at full water depth without further

air drying of the solids typically results in a solids concentration of 6 to 10 percent. Solids concentrations of

20 to 30 percent may be achieved for lime sludge under

the same conditions. Some facilities have achieved solids concentrations above 50 percent by stopping the

influent into the lagoon and allowing drying through

Figure 4-7. Wedgewire drying bed cross section (U.S. EPA,

1979b). Figure 4-8. Dewatering lagoon cross section.

48
evaporation. This process may require well over a year

of holding the solids in the dewatering lagoon.

The lagoon process may incorporate certain modifications similar to sand and/or solar drying bed systems.

Adaptation of a freeze-thaw process to lagoons is common in northern climates.

4.2.4.3 Mechanical Dewatering Equipment

Belt Filter Presses

Belt filter press design is based on a very simple concept. Sludge sandwiched between two porous belts is

passed over and under rollers of various diameters. As

the roller diameter decreases, pressure is exerted on the

sludge, squeezing out water. Although many different

belt filter press designs are used, they all incorporate the

same basic features—a polymer conditioning zone, a

gravity drainage zone, a low pressure zone, and a high

pressure zone.

The polymer conditioning zone can be either a small

tank with a variable speed mixer (approximately 70 to

100 gallons) located 2 to 3 feet from the press, a rotating

drum attached to the top of the press, or an in-line

injector. Press manufacturers usually supply the polymer conditioning unit with the belt filter press.

The gravity drainage zone is a flat or slightly inclined belt

that is unique to each press model. In this zone, solids

are dewatered by the gravity drainage of the free water.

If the solids do not drain well in this zone, problems such

as solids extruding from between the belts and binding

the belt mesh can occur.

The low pressure zone, also called the wedge zone by

some manufacturers, is the area where the upper and

lower belts come together with the solids in between,

thus forming the solids “sandwich.” The low pressure

zone prepares the solids by forming a firm cake able to

withstand the forces within the high pressure zone.

In the high pressure zone, forces are exerted on the

solids by the movement of the upper and lower belts as

they go over and under a series of rollers of decreasing

diameters. Some manufacturers have an independent

high pressure zone that uses belts or hydraulic rams to

further increase the pressure on the solids, thus producing a drier cake.

Belt filter presses can be used to dewater the residuals

produced from either lime softening processes or alum

coagulation. Performance, however, can be affected by

many variables, including solids type and characteristics, conditioning requirements, pressure requirements, and belt speed, tension, type, and mesh.

The type and characteristics of the solids to be dewatered are very important in determining the effectiveness

of belt press dewatering. Lime softening residuals dewater very readily and are efficiently dewatered on belt

filter presses. Since these residuals are more granular

in nature, they can withstand high pressures and easily

dewater to 50 to 60 percent solids.

Conversely, alum residuals are more difficult to dewater

because of the gelatinous nature of the solids. The

dewatering results are variable, depending on the

source of the water coagulated with alum. An almost

pure alum residual is the most difficult to dewater and

must be dewatered at low pressures. The pure alum

residual will dewater to 15 to 20 percent solids. If the

source of the water is a river, and silt and sand are mixed

in with the aluminum hydroxide, the slurry will be more

easily dewatered, with the resulting cake solids between

40 and 50 percent. Even this type of alum residual,

though, must be dewatered at low pressure.

To ensure optimum performance on a belt filter press,

lime and alum solids must first be conditioned with

polymer. Polymer produces a large, strong floc that

allows free water to drain easily from the solids in the

gravity drainage zone of the belt filter press. A typical

belt filter press is shown in Figure 4-9.

Centrifuges

Centrifugal dewatering of solids is a process that uses

the force developed by fast rotation of a cylindrical bowl

to separate solids from liquids. When a mixture of solids

and water enters the centrifuge, it is forced against the

bowl’s interior walls, forming a pool of liquid that separates into two distinct layers. The solid cake and the

liquid centrate are then separately discharged from the

unit. Both types of centrifuges used to dewater WTP

residuals—basket and solid-bowl—use these basic

principles. They are differentiated by method of solids

feed, magnitude of applied centrifugal force, cost, and

performance.

Although commonly used in the past, basket centrifuges

are now rarely used to dewater WTP residuals because

they are a batch process, are more difficult to operate,

and do not perform as well as solid-bowl centrifuges.

The solid-bowl centrifuge, also known as the decanter,

conveyor, or scroll centrifuge, is characterized by a rotating cylindrical conical bowl (Figure 4-10). A helical

screw conveyor fits inside the bowl with a small clearance between its outer edge and the inner surface of the

bowl. The conveyor rotates at a lower or higher speed

than that at which the bowl is rotating. This difference in

revolutions per minute (rpm) between the bowl and

scroll is known as the differential speed and causes the

solids to be conveyed from the zone of the stationary

feed pipe, where the sludge enters, to the dewatering

beach, where the sludge is discharged. The scroll

pushes the collected solids along the bowl wall and up

49
the dewatering beach at the tapered end of the bowl for

final dewatering and discharge.

The differential speed between the bowl and conveyor

is maintained by several methods. Earlier designs used

a double output gear box that imparted different speeds

as a function of the gear ratio. It was possible to vary

the output ratio by driving two separate input shafts.

Eddy current brakes are also used to control the differential. The latest designs can maximize solids concentration through automatic speed control as a function of

conveyor torque.

The solid-bowl centrifuge operates in one of two modes:

counter-current or continuous concurrent. The major

differences in design are the location of the sludge feed

ports, the removal of centrate, and the internal flow

patterns of the liquid and solid phases.

For the most part, solid-bowl centrifuges use organic

polyelectrolytes for flocculating purposes. Polymer use

Figure 4-9. Belt filter press (Andritz Ruthner, 1994).

Cover

Main Drive Sheave

Feed Pipes 


(Sludge and 


Chemical

Bearing 

Base Not Shown

Sludge Cake


Discharge

Centrate


Discharge

Rotating


Conveyor

Differential Speed
 Rotating Bowl

Gear Box

Figure 4-10. Solid-bowl-type centrifuge schematic (Cornwell et al., 1987).

50
improves centrate clarity, increases capacity, often improves the conveying characteristics of the solids being

discharged, and often increases cake dryness. Cake

solids concentrations vary considerably, depending on

the type of alum residuals being dewatered and the

source of the water. High turbidity water yields much

higher cake solids concentrations than does low turbidity water. Lime residuals with high cake solids concentrations dewater very easily. Polymer dosages also vary,

depending on the source of water and the magnesium

and calcium concentrations.

The machine variables that affect performance include:

• Bowl diameter

• Scroll rotational speed

• Bowl length

• Scroll pitch

• Bowl flotational speed

• Feed point of sludge

• Beach angle

• Feed point of chemicals

• Pool depth

• Condition of scroll blades

Many of these variables are preset by the manufacturer,

although some can be adjusted by the operator. A typical

solid-bowl centrifuge is pictured in Figure 4-11.

Pressure Filters

Filter presses for dewatering were first developed for

industrial applications and, until development of diaphragm presses, were only slightly modified for municipal applications. The original models of the press were

sometimes called plate and frame filters because they

consisted of alternating frames and plates on which filter

media rested or were secured. There are currently 20

U.S. installations of filter presses being used to dewater

residuals.

The equipment commonly used to dewater WTP residuals is either the fixed-volume recessed plate filter or the

diaphragm filter press. The diaphragm filter press was

introduced within the last 10 years.

A recessed plate filter press consists of a series of

plates, each with a recessed section that forms the void

into which the solids are pumped for dewatering. Filter

media are placed against each wall and retain the solids

while allowing passage of the filtrate.

The surface under the filter media is specifically designed to ease the passage of the filtrate while holding

the filter cloth. Solids are pumped with high-pressure

pumps into the space between the two plates and individual pieces of filter media. The filtrate passes through

the cake and filter media and out of the press through

special ports on the filtrate side of the media.

The pumping of solids into the press continues up to a

given pressure. When solids and water fill the space

between the filter cloths and no further filtrate flow occurs, pumping is stopped. The press is then opened

mechanically and the cake is removed.

The diaphragm filter press is a machine that combines

the high pressure pumping of the recessed plate filter

press with the capability of varying the volume of the

press chamber. A flexible diaphragm is used to compress the cake held within the chamber. A two-step

process is used, with the compression of the diaphragm

taking place after the initial pumping stage. The release

of water at low pressures helps maintain the integrity of

the floc. After water release appears complete following

the initial filling period, pumping is stopped and the

diaphragm cycle is initiated. The diaphragm pressure is

applied, using either air or water on the reverse side of

the diaphragm. Pressures of up to 200 to 250 psi (1,380

to 1,730 kPa) are applied at this stage for additional

dewatering.

When dewatering alum residuals, lime is added as a

conditioning agent. Cake solids ranging from 30 to 60

percent can be achieved, depending on the source of

the alum residual. Lime softening residuals do not need

any conditioning and can be dewatered to 50 to 70

percent solids.

Precoat is generally not needed when inorganic conditioning chemicals, particularly lime, are used. Precoat is

normally used in cases where particle size is extremely

small, or when considerable variability in filterability and

substantial loss of fine solids to and through the filter

media are anticipated. When substantial quantities of

lime are used, cloth washing may require both an acid

and a water wash. A medium is needed, therefore, that

is resistant to both acid and alkaline environments.

Vacuum Filters

Vacuum filtration was the most common means of mechanically dewatering WTP residuals until the mid1970s. It has been used to some extent in the water

treatment industry to dewater lime residuals. Alum solids

have not been successfully dewatered on vacuum filters.

A vacuum filter consists of a horizontal cylindrical drum

(Figure 4-12) which rotates while partially submerged in

a vat of solids slurry. The filter drum is partitioned into

several compartments or sections. Each compartment

is connected to a rotary valve by a pipe. Bridge blocks

in the valve divide the drum compartments into three

zones, which are referred to as the cake formation zone,

the cake drying zone, and the cake discharge zone. The

filter drum is submerged to roughly 20 to 35 percent of

51
its depth in a vat of solids; this submerged zone is the

cake formation zone. Vacuum applied to this submerged

zone causes the filtrate to pass through the media,

retaining solids particles on the media. As the drum

rotates, each section is successively carried through the

cake formation zone to the cake drying zone. This latter

zone begins when the filter drum emerges from the

sludge vat. It represents from 40 to 60 percent of the

drum surface and ends at the point where the internal

vacuum is shut off. At this point, the sludge cake and

drum section enter the cake discharge zone, where the

sludge cake is removed from the media. Figure 4-12

illustrates the various operating zones encountered during a complete revolution of the drum.

No conditioning is required when dewatering lime residuals. As with all types of mechanical dewatering

equipment, optimum performance depends on the

type of solids and how the filter is operated. Selection

of vacuum level, degree of drum submergence, types

of media, and cycle times are all critical to optimum

performance.

4.2.5 Drying

The drying of dewatered WTP residuals has historically

revolved around the economics of reducing transportation and disposal costs by reducing solids volume and

water content. Drying to solids concentrations greater

than 35 percent is becoming a regulatory issue in many

areas. For instance, the State of California requires that

the solids concentration of a WTP waste be at least 50

percent before disposal in a landfill. Similar to the dewatering process, the drying process may be carried out

through either open air means or through mechanical

devices.

4.2.5.1 Open Air Drying

Any of the solar drying or lagoon procedures may be

applied to the drying process. Drying depends on the

evaporation mechanism. An extended drying process

may require years to achieve the desired solids concentrations, although various innovations have been used

with extended drying to accelerate the drying process.

One such device uses specially mounted tractors to

furrow and mix the solids to increase their exposure to

sun and air.

4.2.5.2 Mechanical Drying

Of the mechanical drying techniques presented in the

discussion of dewatering only, the filter press has shown

the ability to consistently produce solids concentrations

Figure 4-11. Solid-bowl-type centrifuge (Alfa-Laval Sharples, 1994).

52
greater than 35 percent. The ability of this process to

achieve solids concentrations greater than 50 percent,

as would be necessary in California, is not proven.

Thermal drying of solids from WTPs has not been practiced on a full-scale operation basis in the United States.

Cornwell and Koppers (1990) identify the potential for

using steam-operated dryers to raise the solids concentration of a dewatered metal hydroxide sludge to the 65

to 75 percent range, but this process is untried in fullscale operation. Although many thermal processes have

been used with wastewater solids, including the Best

Process, the Carver Greenfield Process, and various

forms of incineration, they are more applicable to a solid

with a high organic content. No conclusive information

is available to evaluate the effectiveness of mixing

wastewater and water solids together before feeding

into a thermal process.

4.2.6 Additional Residuals Handling

Processes

4.2.6.1 Conveyance

Movement of a solids stream from one process to either

another process or to a disposal point may be accomplished by gravity through a pipeline, pumping through

a pipeline, transporting along a mechanical conveyor, or

transporting by vehicle. The type of conveyance used

depends on the form and concentration of the solids

stream and on the transport distance.

4.2.6.2 Equalization Basins

Equalization basins even out the flow of waste streams.

Equalization takes place to prevent surges of water from

being reintroduced at the head of the treatment plant or

at inlets to other residuals handling processes. Equalization basins can also be used to equalize peak daily,

weekly, or monthly loads.

4.3 Residuals Handling Process

Performance

4.3.1 Process Performance

A preliminary screening of processes may be based on

the percent concentrations that the processes can

achieve. This section provides typical ranges of values

for percent solids that each thickening and dewatering

process may generate; insufficient information exists for

drying processes. Many factors can influence the performance of the process: the type of solids, characteristic of solids, solids concentration of the influent,

climatological conditions for open air processes, variations in influent flow rates, and type and dosage of

chemical conditioner. The values presented here should

only be used for screening purposes. The following

value ranges are used to suggest whether thickening,

dewatering, or drying processes should be used to process metal hydroxide solids:

Desired Residual

Metal Hydroxide

Solids Concentration Suggested Process

≤8% Thickening

8–35% Dewatering

>35% Drying

These processes will generally yield a higher percent

solids concentration for a lime sludge than for a metal

hydroxide sludge. For example, a centrifuge may only

dewater a metal hydroxide sludge to a 15 percent solids

concentration, while the lime sludge is dewatered to 50

percent. When selecting a handling process for a lime

softening WTP, the following concentrations apply:

Figure 4-12. Vacuum filter (U.S. EPA, 1979b).

53
Desired Residual

Lime Solids

Concentration Suggested Process

≤30% Thickening

30–60% Dewatering

>60% Drying

4.3.2 Comparison of Thickening Processes

A comparison of the various thickening processes, including solids loading on the units and the solids concentration of various types of residuals, is shown in

Table 4-2.

4.3.3 Comparison of Dewatering Processes

A comparison of dewatering process performance is

shown in Table 4-3.

4.4 Developing Preliminary Residuals

Processing Alternatives

Many unit process combinations are capable of meeting

any set of residuals processing requirements. Most

facilities need to consider the addition of residuals processing, however, because of increased regulatory requirements, changes in finished water-quality goals, increased external disposal costs, and community

relations issues. Two technical factors are used to determine the residuals processing requirements for a utility. The most definitive is the solids concentration of the

residual flow stream. The second factor is the solids

concentration required for different types of residuals

disposal strategies.

A water utility must identify preliminary combinations of

unit processes that can achieve generic residual processing requirements according to a limited set of selection parameters. The approach needed to achieve the

goal includes these steps:

1. Define the fundamental information needed, attempting to identify a preliminary residual processing alternative.

2. Define a methodology of using the fundamental information to make a preliminary selection of a residuals processing alternative that has the potential of

achieving the desired needs.

3. Present a Preliminary Residuals Processing Selection Matrix that can be used with other fundamental

information to make a preliminary selection.

4.4.1 Residuals Processing Requirements

To identify the ideal combination of unit processes for a

preliminary residuals processing alternative, the specific

requirements of the residuals processing system must

first be defined. To do this, information must be obtained

on residuals disposal limitations, quantity and quality of

residuals sources, resource recovery potential, and residuals mass balance.

4.4.1.1 Residuals Disposal Limitations

Often, the ultimate method of residuals disposal determines the limitations, which in turn define the process

design requirements. The percent solids content of a

residual is the primary criterion used to define the acceptable limits of a disposal option. Of course, landfill

and land application disposal options can be accomplished over a wide range of solids content, but both

disposal options have different equipment requirements

at the low and high solids limits of their ranges. The

following tasks will help establish the disposal limitations

that a water utility must consider for further evaluation

of residual process alternatives:

• Identify available residuals disposal alternatives: The

six most common methods of WTP residuals disposal

used in the water industry are:

– Land application

– Landfilling (monofill)

Table 4-3. Comparison of Dewatering Processes (Cornwell et

al., 1987)

Solids Concentration (%)

Process Lime Sludge

Coagulant

Sludge

Gravity thickening 15–30 3–4

Scroll centrifuge 55–65 20–30

Belt filter press 10–15 20–25

Vacuum filter 45–65 25–35

Pressure filter 55–70 35–45

Diaphragm filter press N/A 30–40

Sand drying beds 50 20–25

Storage lagoons 50–60 7–15

Table 4-2. Comparison of Thickening Processes (Cornwell et

al., 1987; Cornwell and Koppers, 1990; Brown and

Caldwell, 1990)

Process Residual

Solids

Loading

Solids

Concentration 

(%)

Gravity Carbonate 30 lb/day/ft2 15–30

Gravity Hydroxide 4.0 lb/day/ft2 1–3

Flotation Hydroxide 20 lb/day/ft2 2–4

Gravity belt Hydroxide N/Aa 2.5–4.5

a No solids loading rate is shown for the gravity belt thickener as it is

not comparable to the values for the gravity thickener and the DAF

unit. Care must be taken in the use of solids loading and percent

solids values for both the flotation and gravity belt thickening, due

to the absence of operating experience for those processes.

54
– Direct stream discharge

– Landfilling (co-disposal)

– Discharge to sewer

– Residual reuse

• Identify disposal limitations for each alternative: The

normal range of acceptable residual solids content

for the six common methods of disposal are:

– Land application, <1% to 15% solids

– Landfilling (co-disposal), <15 to 25% solids

– Landfilling (monofill), >25% solids

– Discharge to sewer, <1% to 8% solids

– Direct stream discharge, <1% to 8% solids (Note:

Allowable solids concentration for this disposal

method may be lower than 8 percent due to regulatory requirements or sewerage concerns such as

deposition.)

– Residual reuse, <1% to >25% solids

• Identify external disposal costs, if they apply: Disposal in a co-disposal municipal landfill may involve

a per unit weight tipping fee levied by an agency

external to the water utility. A user charge can be

assessed for discharges to a sewer, based on flow

rate units and/or the combination flow rate and suspended solids concentrations. In some instances of

residuals reuse, the water utility pays a fee to the

user to cover the additional costs of special handling

or the difference in production costs associated with

accommodating sludge over a more traditional material.

4.4.1.2 Quantity and Quality of Residuals

Sources

For many water utilities, monitoring the quantity and

quality of residuals sources either has not been done,

or has not achieved the same level of accuracy or

frequency as has monitoring of the finished water. This

deficiency may require that additional effort and time be

spent to collect the following information:

• Residuals sources: The typical residuals sources associated with conventional WTPs are:

– Headworks/bar screens

– Presedimentation basins

– Filter backwash

– Grit basins

– Sedimentation basins

– Filter-to-waste

• Flow rates and quantities of residuals sources: Flow

rate data collected for each residuals source should

include the following:

– Maximum flow rate

– Average flow rate

– Maximum volume per release event

– Frequency of events

– Maximum number of events per day

– Potential for concurrent events

– Minimum flow rate

– Average volume of release

– Minimum volume per event

– Average event duration

• Quality of each residuals source: The primary quality

parameter to monitor is percent solids which, in analytical terms, represents total suspended solids. Other

parameters that should be analyzed on a limited basis are:

– Aluminum and/or iron

– Total dissolved solids

– Trihalomethane formation potential

– Trace metals

– Toxicity Characteristic Leachate Procedure (TCLP)

– pH

– Total organic carbon

4.4.1.3 Resource Recovery Potential

Although standard water conservation strategies support the recovery of reusable residuals, the value of

recycle water is being challenged by the potential risks

of recycling infectious organisms, heavy metals, manganese, and DBPs. Typical recycling of untreated residuals streams (e.g., backwash and filter-to-waste

residuals) to the plant influent may result in the accumulation of these contaminants in the treatment processes.

Little documentation as to the extent of this is currently

available. Because of these concerns, the value of water

continues to increase with time. The recovery potential

for reusable residuals streams should be evaluated using the following procedure:

• Assess the value of recovered water: Essentially, the

value of recovered water depends on the value of

raw water and the relative abundance of supply. Establish whether the cost of recovery is justified.

• Identify residuals sources easily recovered: The

backwash and filter-to-waste residuals streams generally comprise the largest volume of residuals generated with the lowest percent solids concentration.

Treatment of these residual streams is usually justified for the value of the water recovered. Other residuals streams need to be evaluated closely for their

potential effect on finished water quality.

• Identify residuals sources unsuitable for recovery:

The most difficult residuals streams to justify are the

55
sidestreams of the mechanical dewatering processes, which usually contain many components that

are not desirable to return to the plant.

4.4.1.4 Residuals Mass Balance

The most valuable exercise in evaluating the residuals

generated at a WTP is the development of a mass

balance. A mass balance is a theoretical accounting of

the water and solids content that come into a plant;

solids added to the water; and the water and solids

content that come out of the plant in the finished water

and residuals streams. This assessment must include

the effects that normal variations in raw water quality

and associated chemical feed rates have on residuals

production rates. The objective of the mass balance is

to specifically define the following parameters, which

become more useful in further evaluations:

• Maximum, minimum, and average solids production

rates for the raw water and each residuals side

stream in terms of percent solids, flow rates, and total

solids produced per million gallons of water treated.

• Allowable return residuals stream limits in terms of

hydraulic and percent solids limit (e.g., maximum

backwash return to plant influent should not exceed

5 percent of operating production capacity).

• Allowable solids content in the finished water.

• Allowable flow rate and solids content for available

residuals disposal options.

• Required residuals processing efficiency for each residual stream (based on the allowable return limits,

effluent limits, and disposal limits for each residual

stream).

The product of this exercise is the identification of potential solids content objectives for each residual stream

process.

4.4.2 Preliminary Selection of Residuals

Processing Alternatives

The next step in developing a preliminary residuals

processing alternative is to select a unit process or

combination of unit processes that can fulfill the defined

processing requirements. There are four general types

of criteria on which the selection process is based.

These are:

• Specific problem-based criteria

• Residuals disposal-based criteria

• Solids concentration-based criteria

• Other selection criteria

In the event that the problem-specific approach does not

apply, the secondary level of selective review addressed

the available disposal methodology. The third selective

review focuses on the residuals solids concentration

categories.

The Preliminary Residuals Processing Selection Matrix

shown in Table 4-4 provides a finite set of unit process

combinations categorized according to three ranges of

residual solids concentrations that they would be used

to treat.

The rules of selection generally follow the basic premise

that most water utilities will want to minimize the capital

and labor costs of residuals processing. The preliminary

selection should attempt to find the lowest solids content, operationally simple, single-unit process that can

fulfill the requirements. The preferred alternatives, from

best to worst, are:

• Low solids; simple operations; single unit process.

• Low solids; simple operations; combined unit process.

• Low solids; complex operations; single unit process.

• Low solids; complex operations; combined unit process.

• Medium solids; simple operations; single unit process.

• Medium solids; simple operations; combined unit

process.

• Medium solids; complex operations; single unit process.

• Medium solids; complex operations; combined unit

process.

• High solids; simple operations; single unit process.

• High solids; simple operations; combined unit process.

• High solids; complex operations; single unit process.

• High solids; complex operations; combined unit

process.

The following approach can be used to select a preliminary residuals processing alternative that can be

evaluated in detail. Site-specific factors such as land

availability, climate, community, and environmental concerns will play a significant role in the selection process.

4.4.2.1 Specific Problem-Based Selection

At the first level of selective review a WTP operation

must consider the following residuals processing issues:

• Regulatory related issues: These issues result from

the ever-changing regulatory environment. Typical

concerns are land application and landfill disposal

regulations.

• Finished water-quality issues: With the recent publication of the draft Disinfection/Disinfection Byproduct

Rule (D/DBP), the most prevalent concern facing the

water industry is how to deal with the residuals byproducts of enhanced coagulation and the increased

56
Table 4-4. Preliminary Residuals Processing Selection Matrix

Problem-Specific Alternative Categories

Single Unit Process

Alternativesb Co

Residual SolidsConcentrationCategories

Residuals DisposalAlternativeCategories

Regulatory RelatedIssues

Finished WaterQuality Issues

DisposalEconomic Issues

CommunityRelation Issues

SimpleOperations

ComplexOperations Simple O

Low solids content(<1 to 8%) by dry

weight

Direct discharge• To source water• To water of USA• Intermittent streamSewer disposalLand application• Agricultural• Forest land• Disposal siteResidual reuse• Landscape treenursery• Irrigation of turf grass

NPDES permit for directdischargeWater supplier policychange on return waterquality (i.e., “zerodischarge”)Residual disposalregulations (if applicable)Aquifer protectionrequirements

Increased sludgeproduction due toenhancedcoagulation (EC)Increased volume ofresiduals due toimplementation of“filer-to-waste”Concern forincreased arsenicconcentrationsassociated with ECarsenic removalIncreased concernwith recycling ofinfectious organismsin the recoveredwater

Change in sewerservice charges forsludge disposalRestrictions onsolids content insludge for sewerdisposalLand applicationdisposal requiresnutrient contentaugmentationLoss of market forreuse product

“Not in my backyard” syndrome(NIMBY)Potential odorsConcern with trafficof transport vehiclesConcern for risk ofincidental exposureConcern for aerialdispersion ofbiologicalcontaminantsGeneralenvironmentalconcerns of public

Direct dischargefrom processbasin (DDPB)Lagoon (LAG)Sand drying bed(SDB)Solar drying basinwith decant (DB/D)Wedge wire filterbed (WFB)

Gravity thickener(GT)h

Dissolved airflotation thickener(DAF)hMechanicalthickening (MT)d,hVacuum assistedplate filter bed(VAFB)g,hBackwash watersedimentationbasin (BWSB)c

DDPB + residstation (PS)PS + LAG + station (DPS)PS + DB/D + PS + WFB + PS + SDB + Backwash restation (BRPSEqualization pump station 

Medium solidscontent (<8 to 25%) Land application • Disposal

• Agricultural• ForestResidual reuse• Sod amendment forturf farms• Landscape treenursery• Coagulant recoveryLandfill• Co-disposal withmunicipal solid waste(low moisture contentrefuse)

Residual disposalregulation (if applicable)Residual reuseregulations onagricultural landLandfill operationrequirementsDisposal of hazardouswaste byproducts fromcoagulant recovery

Increased sludgeproduction due to ECPotential forhazardous wasteclassification witharsenicOrganics, color, and heavy metalconcentrations inrecovered coagulantConcern withrecycling ofinfectious organismsin recovered water

Limit on minimumsolids contentIncrease in tippingfee per ton at landfillLoss of reuseproduct marketFuture refusal toaccept sludge atlandfillLand applicationmay require nutrientaugmentation

NIMBYPotential odorsNoise pollution withsludge handlingequipmentTraffic concernsIncident exposureriskEnvironmentalconcerns

LAGSDBDB/DWFB

MTd

VAFBgMechanicalDewatering (MD)e

PS + LAG + PS + SDB + PS + DB/D + PS + WFB + BRPS + LAG 

57
Problem-Specific Alternative Categories

Single Unit Process

Alternativesb Co

Residual SolidsConcentrationCategories

Residuals DisposalAlternativeCategories

Regulatory RelatedIssues

Finished WaterQuality Issues

DisposalEconomic Issues

CommunityRelation Issues

SimpleOperations

ComplexOperations Simple O

High solids content(>25%)

Landfill• Monofill• Use as landfill coverextenderResidual reuse• Coagulant recovery• Integrated intomasonry products• Filler in Portlandcement

Landfill designrequirementsLandfill operationrequirementsDisposal of hazardouswaste byproduct fromcoagulant recoveryResidual reuserequirements (if any)

Increased sludgeproduction due to ECNeed to segregateunrecoverablesidestreams of highsolids unit processesPotential hazardousclassification ofsludge due toarsenic and othermetal concentrations

Future refusal toaccept sludge atlandfill as coverextenderRefusal to acceptfuture sludge inlandfillLoss of reuseproduct market

NIMBYTraffic concernsNoise pollution fromdewateringequipmentLong-termenvironmentalconcerns

LAG + extendeddrying activities(ED)SDB + EDDB/D + EDWFB + ED

VAFP + EDh

MDe

PS + LAG + PS + SDB + PS + DB/D + PS + WFB + PS + LAG + extended dry(EDA)PS + SDB + PS + DB/D + PS + WFB + BRPS + LAG EDA

a Assumes residuals collection systems exist in process basins with gravity flow to residual processing and excludes sludge storage, transport, and disposal related equipment. b Assumes residuals collection systems exist in process basins with gravity flow to residuals pump station and excludes sludge storage, transport, and disposal equipment.c These alternatives are additional unit processes necessary to accommodate backwash flow rates. d Mechanical thickening includes gravity belt thickeners, centrifuges, etc.e Mechanical dewatering includes belt filter presses, centrifuges, filter presses, etc., plus cake handling equipment. f This combination offers good potential for optimum recovery of used water.g This combination is commonly used for coagulant recovery. h These unit processes include thickened sludge pump stations.

Key

BRPS = Backwash recovery pump station

BWSB = Backwash water sedimentation basin

DAF = Dissolved air flotation

DB/D = Solar drying basin with decant

DDPB = Direct discharge from process basin

DPS = Decant pump station

EB/PS = Equalization basin and pump station

EC = Enhanced coagulationED = Extended drying

EDA = Extended drying area

GT = Gravity thickening

LAG = Lagoon

MD = Mechanical dewatering

MT = Mechanical thickening

PS = Residuals pump station

SDB = Sand drying bed

VAFB = Vacuum-assisted plate filter bed

WFB = Wedge wire filter bed

Table 4-4. Preliminary Residuals Processing Selection Matrix (Continued)

58
volume of residuals that will be processed and disposed of. This category also addresses concerns associated with returning untreated residuals to the

treatment process.

• Disposal economics issues: These issues require existing residuals processing practices to be revisited

in response to cost increases for disposal activities

that are interutility and external to the water utility.

• Community relations issues.

Table 4-4 provides a brief description of these issues

and identifies residuals disposal alternatives associated

with low-, medium-, and high-solids content residuals.

To use the matrix for the selective review process, a

utility should first identify the Residuals Solids Concentration category (low, medium, high) that applies to its

current capabilities. Next, the utility should identify,

within its concentration category, a specific problem that

is similar to one currently facing the utility. After these

two steps, a preliminary selection can be made from the

four right-hand columns, using the priority protocol listed

in the previous section. Ideally, the selection made will

use the maximum number of existing unit processes.

This approach can be rationalized by the obvious economic benefit of reusing existing facilities in a process

group to meet the processing requirements.

4.4.2.2 Residuals Disposal-Based Selection

If a relevant problem cannot be identified on Table 4-4,

the utility should select a preliminary alternative based

on the available solids disposal option with the highest

probability of being implemented. Selection of the preliminary alternative follows the same procedure described in the previous section, using the Residuals

Disposal Alternatives column instead of the ProblemSpecific one. The scope of each disposal alternative is

outlined below:

• Landfilling: The four most common landfill disposal

techniques for WTP residuals are:

– Co-disposal with municipal solid waste (refuse)

– Use as daily landfill cover extender

– Monofill of WTP residuals only

– Co-disposal with WWTP biosolids

• Land application: The three most common land application techniques for WTP residuals are:

– Land application on agricultural land

– Land application on forest land

– Land application on a designated disposal site

• Sewer disposal: This alternative assumes that the

sewer ends at an existing WWTP.

• Direct discharge to source stream: Several configurations of direct discharge are possible:

– Discharge to waters of the United States.

– Discharge to a water district supply canal.

– Discharge to an intermittent stream bed.

– Discharge to a dry arroyo (a watercourse in the

arid southwest that only typically flows during, or

directly after, a storm event).

• Residuals reuse: Although WTP residuals do not

have the inherent fertilizer value of WWTP residuals,

these reuse alternatives are being practiced:

– Recovery of coagulants

– Recovery of lime at softening plants

– Use in landscape tree nursery (low solids)

– Use in making bricks (high solids)

– Use in Portland cement (high solids)

– Use in turf farming

4.4.2.3 Solids Concentration-Based Selection

The most basic level of review is to identify preliminary

processes that can produce a specific range of solids

concentrations. Three possible categories of solids concentration are:

• Low: Low processes can produce solids concentrations ranging from less than 1 percent to 8 percent

solids, by weight. This group represents the general

limits of the thickening processes, specifically the

common gravity thickener.

• Medium: Medium processes can produce solids concentrations ranging from greater than 8 percent to 25

percent solids, by weight. This group represents the

general limits of the simple dewatering processes,

such as sand drying beds, and general mechanical

dewatering with belt filter presses and general service centrifuges.

• High: High processes can produce solids concentrations of greater than 25 percent solids, by weight.

These are the general limits of the high solids dewatering processes, such as filter presses and high-solids centrifuges. These limits are also associated with

the minimum solids content for a monofill, although

50 percent solids is preferable.

4.4.2.4 Other Selection Criteria

The matrix also identifies unit processes and combined

unit processes used to accomplish optimal water resource recovery alternatives and coagulant recovery

alternatives.

59
4.5 Specific Residuals Unit Process

Selection Criteria

Selection of residuals handling processes is a more

complicated task than selection of traditional drinking

water treatment processes for the following reasons:

• Little operating experience with residuals handling

processes exists for use as a basis of comparison.

• Residuals handling processes are more difficult to

test with a procedure such as a jar test or a pilot filter

because of the interdependence of solids handling

on other unit processes (e.g., coagulation, sedimentation, and thickening).

Five selection criteria may be used in screening and

selecting residuals handling processes:

• Discharge limitations and the effective operating

range of the residuals handling process.

• Similar operating experience with unit process.

• Bench- and pilot-scale testing of unit process.

• Construction, operation, and maintenance costs.

• Environmental impact of unit process.

Each of these selection criteria is described in more

detail in this chapter. The flowchart in Figure 4-13 shows

how the five selection criteria (in boxes) are used to

identify a desired process alternative. No attempt is

made to weigh the relative impacts of these criteria,

which vary from facility to facility.

4.5.1 Operating Experience

The effectiveness of residuals handling processes are

very specific to the characteristics of the raw and processed water and, in some cases, to environmental conditions. Bench-, pilot-, and full-scale tests provide

information needed to select processes. Circumstances

may dictate that no testing or very little testing can be

performed by a utility; in this case, the experience of

other utilities with various residuals handling processes

may be the best source of information. The Water Industry Data Base (WIDB), developed by the American

Water Works Research Foundation (AWWARF) and the

American Water Works Association (AWWA), makes

this information available (Kawczyinski and Achterrman,

1991).

This survey covered 438 utilities and 347 WTPs. Treatment methods and disposal practices were categorized

for each facility. The different thickening and watering

categories are summarized in Tables 4-5 and 4-6. Note

that, since the database was published in 1991, a series

of new facilities have started operation using processes

such as vacuum-assisted drying beds. Consult with

equipment manufacturers for more recent operating

histories.

What combination of unit processes


can operate effectively in this range?

What is the similar operating experience


with each alternative?

Screen alternatives

What are the relative environmental


impacts of each alternative?

Is it feasible to perform 


bench- and/or pilot-


scale testing?

Develop and


perform bench/


pilot testing.

Size alternative residuals


handling processes.

Develop and compare


alternative costs.

Select residuals handling process.

YES

NO

Develop preliminary process alternatives.

What are the discharge limits?

Figure 4-13. Residuals handling process selection flow chart.

Table 4-6. Survey of Dewatering Methods at Water Treatment

Plants in the United States (Kawczyinski and

Achterrman, 1991)

Treatment Methods

Existing Number of

Facilities

Lagoons 180

Sand drying beds 26

Freeze assisted drying beds 33

Solar drying beds Number included with sand

drying beds

Belt filter press Number included with filter

presses

Centrifuges 10

Filter presses 20

Vacuum filters 4

Screw press 0

Table 4-5. Survey of Thickening Methods at Water Treatment

Plants in the United States (Kawczyinski and

Achterrman, 1991)

Treatment Methods Existing Number of Facilities

Lagoons 180

Gravity thickening 48

Dissolved air flotation 0

Gravity belt thickeners 0

60
4.5.2 Bench-Scale and Pilot-Scale Tests

4.5.2.1 General

Bench- and pilot-scale tests serve two functions in the

selection of residuals handling processes. First, they

provide additional data for the sizing of the full-scale

equipment. Second, they provide an indication of the

process performance (i.e., the solids concentration that

can be achieved).

Figure 4-14 is a guide to the use of bench- and pilotscale tests for this application. Bench- and/or pilot-scale

testing should be considered after the alternative residual handling schemes have been screened and narrowed down to a few options. Pilot- and bench-scale

testing may not be needed if there is similar experience

with the process at a comparable facility, defined as one

of similar size with approximately the same raw water

quality and unit processes. Short implementation schedules and limited study budgets may curtail bench- and

pilot-scale testing, but the lack of related experience can

make bench- and pilot-scale testing desirable.

4.5.2.2 Bench-Scale Testing

These bench-scale tests are generally used for thickening and dewatering processes (Dentel et al., 1993):

• For thickening: settling tests, flotation test, and capillary suction time test.

• For dewatering: time-to-filter test, filter leaf test, capillary suction time test, and settling tests.

Detailed lists of equipment and test procedures for each

of these tests are available (Cornwell et al., 1987; Dentel

et al., 1993). The performance of a bench-scale test

does not have any direct correlation to the performance

of a given process. Bench-scale tests typically simulate

full-scale operations. Thus, a correlation may be developed, for instance, between the results of capillary suction time tests and the performance of centrifuges. The

application of the bench-scale test is then to analyze the

effect of different sludge conditioners. Bench-scale tests

may have application to the selection of a residuals

handling process if the bench-scale tests can be correlated to process performance based on previous testing.

Equipment suppliers should be asked about the availability of past bench-test comparisons. Suppliers, often

times, can bench test units at their own facilities.

4.5.2.3 Pilot-Scale Testing

Pilot-scale testing involves the use of test equipment

similar enough in size to full-scale operating equipment

that the test results can be directly compared between

pilot-scale test and full-scale operation. This testing may

involve the construction of a scaled-down version of a

process basin such as a gravity thickener or a sand

drying bed. Suppliers of mechanical thickening and dewatering equipment typically have test units available for

rent or loan, and can test either on site or at the supplier’s facilities.

4.5.2.4 Test Procedures and Analysis

The use of bench- and pilot-scale tests has many pitfalls. The test must be as representative as possible of

the real world conditions, and must simulate all of the

processes involved with the alternative in question. This

is particularly a concern with thickening and dewatering,

where the incoming solids must represent the solids that

will actually be concentrated and transported from the

previous process. Even when the bench- or pilot-scale

testing successfully models the performance of a process, the results may not reveal operational problems

that are part of the full-scale process.

The duration of testing is a particular concern. Yearround operation may involve significant changes in raw

water quality that can affect the operation of the residual

handling process. One week of pilot-scale testing may

not reflect this impact, and year-round pilot-scale testing

is usually not feasible. The use of short-term pilot-scale

testing to correlate the results with bench-scale tests

may come close to approximating the results of longterm testing, though on a lesser scale.

Time and Budget Allow


Bench-/Pilot-Scale Testing?

Enough Similar Experience Available


to Estimate Sizing and Performance?

Screen Residuals Handling Alternatives

Data Available to


Correlate Bench Testing and


Full-Scale Performance?

Identify Availability of


Manufacturer Pilot Equipment Select Bench Tests

Identify Elements


of Program

Perform Pilot Testing

Perform Testing and Analyze Results

Yes No

No

Yes

Figure 4-14. Bench/pilot testing decision flow (Dentel et al.,

1993).

61
When analyzing the results, the limitations of the test

setup and test protocol should be considered. If limited

testing is done, less emphasis should be placed on the

results.

4.5.3 Environmental Impacts

Environmental impacts have become a standard consideration for public facilities, and are typically part of this

selection process. Environmental criteria, however, are

subjective, making them difficult to use as a basis for

comparison. The environmental impacts of a WTP residuals handling process can vary greatly, depending on

whether one’s viewpoint is that of an operator, a neighbor, or even a migrating bird! Some relevant environmental considerations for this selection process are:

• Effectiveness in meeting discharge requirements:

The ultimate criterion for the selection of a process

or a combination of processes is whether the process

will reliably and consistently meet the regulatory discharge requirements. This depends on many factors,

including the size of the application and process loading conditions.

• Ground-water quality: Any process that might release

contaminants that could migrate to the ground water

should be approached with caution. Unlined lagoons

and drying beds are typically of greatest concern.

• Noise: Of concern to both plant employees and site

neighbors, noise is typically associated with thickening and dewatering processes that use mechanical

equipment. Open-air processes may also involve

noise from front-end loaders or other devices used

to move and remove solids. Offsite transport of solids

is another solids concern. Mitigation techniques include hearing safety equipment for workers, acoustically sealed buildings for mechanical devices, and

specified operating hours for front-end loaders and

other vehicles.

• Odors: Many operators emphatically say that no

odors are associated with WTP residuals and, frequently, they are correct. With open-air processes

(and sometimes even mechanical systems), however, odors are a possible consequence. The traditional technique used to mitigate odors is an

operational change in the loading and removal of

solids from process units. Chemicals such as caustic

and chlorine may be needed to stabilize and destroy

odor causing organisms in the sludge. Odor control

units may also be required for systems installed indoors.

• Energy use: Energy use varies with both equipment

usage and the cost of removing solids from the site.

Mechanical dewatering techniques require the highest energy consumption, although consumption varies widely between the different processes. Different

applications of processes can also affect the percent

solids concentration of the residual before disposal.

A higher concentration translates to a lower volume

for disposal. This lower volume then results in the

use of less transport energy.

• Insects and other pests: Insects are a potential concern with open-air processes. Submerged lagoons or

gravity thickeners may serve as breeding grounds for

mosquitoes and other flying insects. Flies and gnats

may be a concern when drying residuals in the open

air.

• Impact on the neighborhood: This broad category

encompasses many environmental concerns that can

result in the lowering of local property values.

• Air pollution: A typical concern when evaluating

wastewater processes, air pollution (stripping volatile

contaminants into the air) is a small concern with

water treatment systems. Some concerns may arise

from the potential use of mechanical equipment having gas or diesel engines.

• Space requirement: Many plant sites do not have a

large property area or the ability to expand onto adjacent vacant land. Limited space typically drives

process selection away from open-air processes and

toward mechanical processes.

• Employee and public safety: Mechanical handling

processes typically involve the highest safety concerns because they use heavy equipment with the

potential to cause serious accidents. All processes

involve some potential hazard, however, ranging from

suffering heat stroke while working in a solar drying

bed to drowning in a thickener.

In comparing environmental criteria, the interrelationship of these criteria with both construction and operation and maintenance costs is noteworthy. For example,

the ground-water quality concerns associated with lagoons can be resolved by the installation of a relatively

impermeable liner beneath the lagoons; the liner mitigates the environmental concern at a significant increase in construction cost. The relationship between

environmental controls and operation and maintenance

costs should be considered in the analysis of different

residuals handling approaches.

4.6 Final Screening of Residuals

Handling Processes

4.6.1 Additional Selection Criteria

Previous operating experience, environmental impacts,

and bench- and pilot-scale testing are criteria common

to most residual handling process selections. Additional

criteria applicable to both wastewater processes and

water treatment process are (U.S. EPA, 1979b):

62
• Compatibility: Compatibility with existing land-use

plans; areawide, solid waste, water, and air pollution

controls; and with existing treatment facilities.

• Implementability: Available sites; available personnel

(including specially skilled labor); sufficient or expandable existing utilities; meets regulatory review

requirements; can be achieved within the required

schedule; whether existing processes can be interrupted for construction connection.

• Flexibility: Ability to respond to new technology,

changing regulations; changing loads; incremental

expansion.

• Reliability: Vulnerability to disasters; probable failure

rate; backup requirements; operator attention required.

• Impact of residual handling process sidestreams on

existing process.

4.6.2 Process Alternative Selection Matrix

Process combinations that achieve different degrees of

residuals solids concentrations are shown in Table 4-4.

The final selection of the process combination depends

on the relative importance of the selection criteria. One

approach uses a weighting table to aid a decision-maker

in analyzing different process combinations in terms of

weighted criteria (U.S. EPA, 1979b). An example of this

approach is presented in Table 4-7. The weights presented in this table are examples only; each facility must

assign the appropriate relative weights according to its

own priorities.

4.7 Residuals Handling Process Design

Issues

4.7.1 Mass Balance Diagrams

4.7.1.1 Estimating Solids Production

Solids mass balances are used to develop estimates of

hydraulic and solids loading on residuals handling unit

processes. Solids production from a WTP may be estimated based on the raw water suspended solids removal and the quantity of process chemical added.

Refer to Chapter 3 for equations used to estimate solids

product.

When estimating solids production for the preliminary

design of a facility, the following should be considered:

• Historical trends of raw water suspended solids loading must be analyzed to determine the average day,

maximum day, and peak hour loadings. These values

are typically used in evaluating the following design

issues:

– Average day: Used to analyze annual solids disposal fees and land availability for open-air dewatering and drying processes.

– Maximum day: Used to size thickening, dewatering, and drying processes.

– Peak hour: Used to size piping and pumping. If

equipment such as centrifuges or belt filter presses

are sized based on maximum day loadings, then

the residual handling systems must be designed

with the capacity to absorb the difference between

the peak hour and the maximum day loadings. This

may be done by incorporating equalization basins

into the design. Sedimentation basins and thickeners can also equalize peak hourly, daily, and

weekly flows. An alternate approach is to consider

the ability of a given process to handle short-term

spikes in the solids loading rate.

• Most water treatment facilities record solids loadings

in terms of turbidity (NTU) rather than suspended

solids. Methods of converting turbidity values to suspended solids values are available (Cornwell et al.,

1987). Generally, the ratio of suspended solids to

NTU is 1 to 2.

• Estimating raw water suspended solids loadings is

inherently more difficult for new WTPs than is adding

residuals handling facilities to existing WTPs. The

options are:

– To use the historical records of a WTP that uses

similar processes with a comparable water source.

– To develop a water-quality database through a program of sampling, jar testing, filtration, and analysis. Because a short-term program will represent

conditions only for one period in time, such a program may not reflect long-term water quality at

the site.

4.7.1.2 Developing the Diagram

Examples of a mass balance diagram and calculation

are shown in Figures 4-15A and 4-15B, respectively.

The percent of solids captured for the gravity thickeners

and the filter presses are based on the typical values in

Tables 4-2 and 4-3. These values should be verified in

a bench- or pilot-scale testing program. The diagram

can also provide valuable data regarding the quantity of

chemical conditioner—polymer, in this case—which is

required in the process.

4.7.2 Equipment Sizing

WTP residuals handling systems can vary from simple

sewer/stream disposal systems to elaborate thickening/dewatering/drying complexes. The mass balance

examples in Figures 4-15A and 4-15B are only the

starting point; the next step is the generation of a schematic such as that shown in Figure 4-16, which is based

63
on refined flow rates and individual equipment loadings.

Issues that must be identified at this preliminary design

stage are the number of anticipated shifts, coordination

with main plant operations, and provision for redundant

process trains. Equipment size can also be reduced by

providing adequate equalization facilities. Projected

equipment suppliers also play a key role in this phase

as the specific equipment capabilities and requirements

are identified.

Open air dewatering/drying systems are the most common form of residual handling process. While the process approach and the system equipment are relatively

simple, the actual design is dependent on local climatological conditions, regulatory requirements, and projected use. Detailed information on design details for

these systems exists in the literature (Cornwell et al.,

1987; Cornwell and Koppers, 1990).

4.7.3 Contingency Planning

The Process Design Manual for Sludge Treatment and

Disposal (U.S. EPA, 1979b) presents the contingency

planning design issues that should be incorporated into

a WWTP solids handling facility. These issues are

equally applicable to a water treatment facility.

Contingency planning concerns and possible design

resolutions are listed in Table 4-8.

Table 4-7. Example of Weighting System for Alternative Analysis (U.S. EPA, 1979b)

Alternative 1 Alternative 2 Alternative n

Categories and

Criteria

Relative

Weighta ARb WRc AR WR . . . AR WR

Effectiveness:

Flexibility

Reliability

Sidestream effects

3

5

3

4

3

10 

12

15

30

6

5

9

18

25

27

.

.

.

.

.

.

.

.

.

6

2

7

18

10

21

Compatibility:

With existing land

use plans

2 8 16 8 16 . . . 4 8

With areawide

wastewater, solid

waste, and air

pollution programs

3 3 9 6 18 . . . 7 21

With existing

treatment facilities

4 5 20 5 20 . . . 3 12

Economic impacts:

Construction cost

4 7 28 8 32 . . . 7 28

O&M cost 1 8 8 9 9 . . . 8 8

Total weighted

alternative ratingd - - 1,576 - 1,430 - - - - 1,317

a Relative importance of criteria as perceived by reviewer; scale, 0 to 5; no importance rated zero, most important rated 5.

b Alternative rating. Rates the alternative according to their anticipated performance with respect to the various criteria; scale 1 to 10; least

favorable rated 10. c

 Weighted rating. Relative weight for each criteria multiplied by alternative rating.

d Sum of weighted ratings for each alternative.

Q = 50


SS = 10 Fe Polymer

Q = 0.251 Q = 0.268 S = 1117560

Q = 0.017 S = 27980

S = 12417

Q = 0

Gravity Thickening

Filter Press

Figure 4-15A. Mass balance schematic.

64
Backup systems and storage spaces are key contingency planning items that are often incorporated into

residuals handling facility designs. An aerial view of a

residuals handling facility is shown in Figure 4-17. This

facility uses centrifuges as the primary mode of dewatering along with solar drying beds. The drying beds

serve the following purposes:

• Handle peak solids loadings that the centrifuges may

not be able to effectively treat.

• Serve as a dewatering process in case of centrifuge

failure.

• Provide a dewatered solids storage area in the event

that offsite solids hauling is disrupted.

4.7.4 Specific Design Elements of

Mechanical Dewatering Systems

Mechanical dewatering systems are generally the most

complex of the residuals handling systems. Layout examples for belt filter presses, centrifuges, filter presses,

and vacuum filters are shown in Figures 4-18, 4-19,

4-20, and 4-21.

Figure 4-15B. Mass balance calculation.

65
Figure 4-16. Preliminary residuals handling process schematic (Bratby et al., 1993).

66
These schematics identify common elements of mechanical dewatering system design:

• Redundant dewatering units.

• Chemical conditioner (polymer) feed system.

• Simplified approach for removing dewatered solids

from the unit.

• Byproduct water removal system.

• Mechanism for removing dewatering unit from building for repair.

Additional design items that should be considered are:

• Local control of sludge feed into the units.

• Provisions for equipment washdown and drainage.

• Acoustic paneling in room to dampen noise.

• Odor control system (rarely implemented in water

production facilities).

• Local shower facilities for staff.

Table 4-8. Residual Handling Facility Contingency Planning

Issues (U.S. EPA, 1979b)

Potential

Problems Potential Resolutions

Equipment

breakdowns

• Provide redundant units and piping.

• Allow for solids storage during repair.

• Base design on single shift; second or third

shift can be utilized on an emergency basis.

Solids disposal

disruption

• Plan for a secondary disposal approach.

• Allow for solids storage during disruption.

Solids production

greater than

expected

• Provide for safety factors in process design.

• Allow for solids storage during peak periods.

• Base design on single shift; second or third

shift can be used on emergency basis.

Figure 4-17. Aerial view of residuals handling process system, Val Vista WTP, Cities of Phoenix/Mesa, AZ (John Carollo Engineers, 1994).

67
4.8 Air Emissions Control

Along with traditional residuals generated by the physical/chemical treatment of water, some treatment processes can generate gaseous residuals that can be

released to the atmosphere. Gaseous residuals form

two groups: emissions intentionally generated as a

byproduct of a treatment process, and emissions generated by the accidental release of a gaseous chemical

used in water treatment.

4.8.1 Gaseous Residual Byproducts

The treatment of potable water can generate various

types of gaseous residuals, from the simple release of

hydrogen sulfide caused by aerating sulfur water, to the

exhaust of various volatile and synthetic organic compounds from a stripping tower treating contaminated

ground water. Other gaseous byproducts include the

ozone-tainted air exhausting from an ozone contactor at

a modern WTP, and radon released from contaminated

source water.

4.8.1.1 Treatment of Inorganic Gaseous

Emissions

The release of naturally occurring inorganic gases found

in ground water (i.e., hydrogen sulfide) to the atmosphere is not an environmental concern unless it exceeds

a 1 part per million (ppm) concentration. As urban development continues to encroach upon existing water

treatment facilities, the concentration of gases in the

exhaust from aeration and stripping processes may ultimately need to be reduced for aesthetic reasons,

namely odor control. Aeration is a process that mixes air

and water, normally by injecting air into the bottom of a

tank of water; spraying water into the air; or allowing

water to aerate by cascading over an irregular surface

like an artificial stream bed. Stripping is a highly controlled form of aeration engineered for the optimum removal

of a specific gas at a maximum concentration by spraying water into a variable counter-current flow of air

through a stripping chamber.

Hand Railing Floor Opening

Discharge

Upper Level Plan

Monorail on C of machine

Machine Control Panel

Reactor


Filtrate


Water

Sludge 


Discharge

Filtrate


Return Pump

Sludge met


 sub-system

Polymer met


sub-system

Elevation

Figure 4-18. Belt filter press—example of system layout (U.S.

EPA, 1979a).

Sludge Feed Line

Polymer


Tank

Feed


Pump

Centrifuge

Electrical Controls

Plan View

Polymer Feed Line

Centrate


Return

Sludge

Sludge


Pump

Centrate


Pump

Section View

Figure 4-19. Centrifuge—example of system layout (U.S. EPA,

1979a).

68
In cases where aeration is used, concentrations of hydrogen sulfide gas in the exhaust are below the 1 ppm

threshold and are expelled directly into the atmosphere.

If the gas concentrations in exhaust from an aeration

process need to be controlled, the exhaust must be

collected and delivered to a scrubber. Collection is normally achieved by installing a cover over the aeration

process and using an exhaust fan to deliver the air to a

scrubber system. When stripping is used, the exhaust

gas is likely to exceed the 1 ppm concentration and

would probably be directed to a scrubber system.

In air pollution control, a scrubber is an air cleaning

process in which a contaminant gas is absorbed into a

scrubbing liquid. The most common method of scrubbing is accomplished with a packed tower, a vertical

chamber filled with various types of packing materials

(see Figure 4-22). Process exhaust air is discharged into

the bottom of the chamber and flows upward around the

packing materials while scrubbing liquid trickles down

through the packing in a counter-current flow pattern.

The resultant sinuous pathway for the air and liquid

provides an extended exposure of the air to the liquid

and allows the gas to transfer from a gaseous state into

a solution state and react with the scrubber liquid.

In the scrubbing of hydrogen sulfide, the scrubbing liquid

is typically sodium hydroxide (caustic soda), which ultimately is converted to sodium sulfate through the reaction with hydrogen sulfide. To maintain scrubber

efficiency, the scrubbing liquid is periodically replaced

when it can no longer absorb the target gas. The sodium

sulfate liquid is typically discharged directly into a sewer,

as it is a nonhazardous residual. Other disposal alternatives must be evaluated in the event that a sewer is not

available.

4.8.1.2 Treatment of Volatile Organic Gaseous

Emissions

In many industrialized areas of the world, ground water

has been contaminated by the careless disposal of various manmade organic chemicals. Most of these organic

chemicals are very toxic and must be removed from the

water before it is suitable for potable use. A group of

these chemicals is called volatile organic chemicals or

Sludge to Filter Presses

Press Feed Pumps

Sludge Feed Lines

Wash Acid


Line

Filtrate 


& Acid


Return

Sludge Supply


Line

Lime Supply


Line

Acid Tank

Acid


Pump

Sludge


Conditioner

Control


Panel

Floor


Opening

Filter


Press

Platform

Filter Press

Plan View

Elevation View

Future


Press

Acid


Tank

\Feed Pumps

Sludge


Conditioner

Sludge Transfer


Pumps

Figure 4-20. Filter press—example of system layout (U.S. EPA, 1979a).

69
VOCs (e.g., trichloroethylene, known as TCE) because

they can be easily removed from the water by violent

agitation using a stripping chamber similar to that described above for removing inorganic gases.

Stripping chambers are similar to packed scrubbing towers except they have large blowers that can be regulated

to vary the amount of air being forced into the chamber.

Controlling the air flow rate relative to water flow rate

regulates the stripping of the VOC from the water, which

flows counter-current to the air flow. Depending on the

concentration of the VOC being stripped and the local

air quality standards, the stripper exhaust may be discharged directly into the atmosphere. Under these conditions, the volume of air forced through the stripper can

effectively diffuse the concentration of the organic to

trace levels.

In cases where the stripper has unacceptable concentrations of VOCs in the exhaust, it is discharged into an

afterburner that incinerates the organic constituents. An

afterburner is essentially a combustion chamber with an

external fuel source that heats the air flowing through

the chamber to a temperature sufficient to ignite the

volatile organic gases. The size and temperature requirements of an afterburner depend on the specific

organic chemical and the volume of air into which it is

being vented.

Control


Panel

Vaccum and


Filtrate Assembly

Preccat


Assembly

Sludge 


Conveyor

Vacuum


Filter

Vacuum


Filter

Plan View

Vacuum


Filter

Vacuum and 


Filtrate 


Assembly

Sludge Chute Sludge Pumps


and Mechanical 


Equipment

Elevation View

Sludge 


Conveyor

Figure 4-21. Vacuum filter—example of system layout (U.S. EPA, 1979a).

70
4.8.1.3 Treatment of Nonvolatile Organic

Gaseous Emission

A counterpart to the VOCs found in contaminated

ground water is the nonvolatile organics group. These

organic compounds are more difficult to strip, and their

acceptable concentration in the exhaust is so limited

that stripping with an afterburner is not a viable removal

method. In these cases, the chemicals are removed

using liquid-phase adsorption with granular activated

carbon (GAC). Depending on the organic compound

and the acceptable atmospheric discharge limits, the

GAC can be regenerated or must be disposed of as a

hazardous solid waste, as is the case with GAC containing dibromochloropropane (DBCP). Regeneration of

GAC can be used for some nonvolatile compounds, but

the regeneration exhaust is usually cycled through an

afterburner for neutralization by incineration.

4.8.1.4 Treatment of Ozone-Contaminated Air

The use of ozone in potable water treatment has become more common within the water industry and may

become even more popular as an alternative disinfectant. Although ozone is not the panacea to the pending

Disinfectant/Disinfection Byproduct (D/DBP) Rule, it

may be the only method that can truly deactivate infectious organisms like Cryptosporidium. When used in

potable water treatment, its discharge into the atmosphere must be prevented.

Ozone is typically generated in an air flowstream or a

pure oxygen flowstream discharged at least 22 feet

below the water surface in a completely enclosed ozone

contact chamber. Although the half-life of ozone is very

short, trace concentrations in the basin exhaust are

common. Two of the processes that are commonly used

to eliminate the potential release of ozone into the atmosphere are thermal and catalytic destruction.

Thermal ozone destruction systems rely on the rapid

decay of ozone at elevated temperatures. By heating the

contactor off-gas to approximately 300 to 350 degrees

Celsius for 5 seconds, the ozone is converted back to

oxygen. The air is heated by a heat exchanger supplied

with heat from electrically powered heating coils. As

thermal pollution is also a concern in many places, the

heated exhaust can be passed through an air-to-air heat

exchanger, where heat is transferred to incoming air and

thereby kept out of the atmosphere.

Catalytic ozone destruction units have only been in use

since 1990. Although the true composition of the catalysts is considered proprietary by their manufacturers, in

all designs the ozone is destroyed by electron-level

exchanges and manipulations. Because the catalysts

are subject to severe damage if exposed to moisture,

they are usually combined with an air heater, which can

also be used to regenerate the catalyst. The key advantage of the catalytic process is that it has the lowest

energy costs of the two processes.

4.8.2 Accidental Release of a Gaseous

Treatment Chemical

The oldest and most common water treatment process

is chlorination, which, until recently, had been exclusively accomplished using compressed chlorine gas delivered and stored in cylindrical pressure vessels. The

water industry has learned to handle the toxic chlorine

gas with great care, but accidents do occur, usually

resulting in the release of chlorine gas into the atmosphere.

The potentially deadly impact of the release of large

amounts of chlorine has prompted many changes during

the past 6 years in the applicable codes and laws regulating its use. In response to the increased regulatory

constraints, some water utilities have modified their

treatment process to reduce or eliminate the use of

chlorine, and have implemented systems to provide

neutralization of the chlorine in case of a leak. In some

communities, public awareness and fear have forced

the utilities to use a liquid form of chlorine (sodium

hypochlorite/bleach). The liquid form is safer and essentially has the same disinfection capacity as the gas, but

Off Gases

Liquid Distributor

Packing

Liquid Redistributor

Influent Pipe

Blower Treated Water

Figure 4-22. Packed tower (ASCE/AWWA, 1990).

71
the liquid chlorine is much more expensive to transport

and adds sodium to the finished water.

Since the use of chlorine is expected to continue into the

future, it is reasonable to expect existing water utilities

to construct chlorine containment and neutralization

systems as required by the current building codes. The

code requires that the accidental release of chlorine

must be contained and neutralized using a treatment

system such that the chlorine gas exiting the treatment

exhaust does not exceed 15 ppm. The neutralization of

chlorine requires the use of a scrubber system that

employs caustic soda as the scrubber liquid. A chlorine

scrubber operates essentially the same way as that

used for inorganic gas emissions control, with one exception—the motive force for the extraction of the chlorine from the containment structure commonly uses a

venturi with a caustic soda spray nozzle discharging into

the throat of the venturi. The exhausted caustic soda

becomes sodium hypochlorite (bleach), which can be

disposed of through beneficial reuse as a disinfectant.

Coincidentally, the gaseous chemicals ammonia and

sulfur dioxide are also commonly used in potable water

treatment and can be neutralized by the same system

as is used for chlorine.

72
Chapter 5

Direct Discharge of Water Treatment Plant Residuals to Surface Waters

Water treatment residuals are primarily produced

through the coagulation, filtration, and oxidation of surface waters to remove turbidity, color, bacteria, algae,

organic compounds and, frequently, iron and/or manganese. One type of residual is generated by the aluminum

and iron salts that are commonly used for coagulation.

A second type of residual results from the lime, sodium

hydroxide, and/or soda ash added to reduce calcium

and magnesium and soften the water.

Historically, direct discharge of water treatment effluent

to surface waters has been the most commonly practiced disposal method. A 1984 water utility survey indicated that approximately 50 percent of the total

residuals generated by 429 utilities (at least 548,820

metric tons) were pumped directly to surface waters

(AWWA, 1986). Most of these residuals were alum

sludges. Recently, this disposal method has been questioned because of concerns about possible risk to public

health and aquatic life. 

5.1 Theory

The chemistry of aluminum and iron in water are similar

to one another; however, iron species are less soluble

than aluminum species are over a wider pH range.

Because of the potentially adverse effects of aluminum

in the environment, more information is available on the

discharge of alum residuals to surface waters than is

information on residuals produced from the addition of

iron salts. Consequently, this section will focus on alum

residuals.

In the United States, aluminum in surface waters is

primarily in solid form, with approximately 31 percent of

the aluminum being dissolved (Andelman, 1973). Aluminum is abundant in the environment, yet in surface

waters above pH 4.0, its concentration is normally less

than 0.1 mg/L. Understanding the chemistry of aluminum is important to understanding the impact of aluminum on aquatic organisms.

5.1.1 Chemical Interactions

Aluminum is amphoteric—soluble in acidic and basic

solutions, but quite insoluble at neutrality. The trivalent

state, Al+3, is the only naturally occurring oxidation state

found in solutions and solids. The aqueous chemistry of

aluminum is extremely complex; the molecular form, or

species and concentration of each species depend upon

pH, complexes with ligands, and, to a lesser extent, on

temperature and reaction time (Campbell et al., 1983).

Aluminum sulfate, used as the primary coagulant in

most water treatment systems, dissociates in aqueous

solutions, and Al+3 bonds with water molecules, hydroxide ions, other inorganic ions, and organic ions or molecules. When mixed thoroughly with turbid waters at pH

levels from 4.0 to 8.5, new compounds form, especially

with phosphates and organics, which are very insoluble

and which, therefore, precipitate (U.S. EPA, 1988a; Driscoll and Schecher, 1988). Other aluminum coordination

compounds are very soluble and only precipitate upon

oversaturation.

While temperature and inorganic and organic ligand

concentration are important factors affecting aluminum

solubility, aluminum solubility is largely controlled by pH

(Baker, 1982; Driscoll, 1985). The optimum pH for coagulation using aluminum salts generally ranges from

4.6 to 8.0. The presence of anions such as sulfate

affects the solubility of aluminum salts and can move the

optimum pH for coagulation to 4.0 (Hundt, 1986). Furthermore, the use of other chemicals in the water treatment scheme will affect the solubility and speciation of

aluminum. Fluoride addition after alum coagulation

causes increased aluminum solubility and an increase

in inorganic Al+3 species concentration (Driscoll et al.,

1987).

The source of aluminum ions in an aqueous environment is important. An additive can affect parameters

important to solubility, such as pH and temperature. For

example, if a loosely bound aluminum ion enters natural

water from a sludge or desorbs from an organic clay

particle, the resulting solubility of aluminum in that water

will change as an indirect effect of the additive (Driscoll,

1985). Speciation is also dependent on the aluminum

source, as it is influenced by the binding capability of

ligands with aluminum ions. Tightly bound crystals of

aluminum hydroxides will solubilize slowly over a period

of months (Smith and Hem, 1972). Aluminum also may

be held tightly within an organic complex and not solubilize readily.

73
The hydrated trivalent ion is the predominant form of

aluminum at pH levels below 4.0. Between pH 5 and 6,

the predominant hydrolysis products are Al(OH)2+ and

Al(OH)2+, while the solid Al(OH)3 is most prevalent between pH 5.2 and 8.8 (Martell and Motekaitis, 1989).

The soluble species Al(OH)4- is the predominant species above pH 9, and is the only species present above

pH 10 (Martell and Motekaitis, 1989). Polymeric aluminum hydroxides appear between pH 4.7 and 10.5, and

increase in size until they are transformed into colloidal

particles of amorphous Al(OH)3, which crystallizes to

gibbsite in acidic waters (Brusewitz, 1984). Within the

pH range of 5 to 6, aluminum bonds with phosphate and

the resulting compound is removed from solution. Because phosphate is a necessary nutrient in ecological

systems, this immobilization of both aluminum and

phosphate may result in depleted nutrient states in surface waters (Brusewitz, 1984). The concentrations of

aluminum in water vary with pH levels and the humicderived acid content of the water. Even at neutral pH

levels, higher aluminum levels have been found in lakes

with a high humic acid content (Brusewitz, 1984; DHHS,

1990).

5.1.2 Toxicity

A key concern regarding the direct discharge of aluminum residuals to waterways is aluminum toxicity in the

aquatic environment. When aluminum is mobilized in

lakes and streams, it may be toxic to aquatic life. It is

difficult to generalize about the environmental impacts

of water treatment residuals discharges on receiving

waterways, because such impacts are inherently dependent on an array of physical, chemical, and biological stream parameters. The characteristics of each case

must be individually assessed to evaluate the potential

threat of water treatment plant (WTP) residuals on the

water and sediment qualities of receiving streams.

The aluminum species concentration causing toxicity

depends on water chemistry, the organism being affected, and the effect monitored. The aqueous chemistry

of aluminum has been difficult to study for several reasons. First, its form and, therefore, its effects vary with

pH. Second, aluminum ions tend to undergo complex

hydrolytic reactions with many ligands of differing chemistries, forming salts of differing solubilities. Additionally,

aluminum is abundant in the environment, making it

difficult to isolate the effects of residuals aluminum from

those of ambient aluminum. Lastly, aluminum has varying solubility rates (i.e., weak complexes dissolve

quickly and strongly bound complexes dissolve slowly)

(George et al., 1991).

5.1.2.1 Aquatic Biota

Burrows (1977) presents toxicity data for several freshwater species of fish, invertebrates, bacteria, and algae.

Studies have shown that under low pH conditions (pH

less than 6), inorganic aluminum can be toxic to aquatic

organisms. Major fish kills in southern Norway were

attributed to the release of aluminum from soil by acid

precipitation (Muniz and Leivestad, 1980). Furthermore,

inorganic aluminum concentrations less than 400 μg/L

had significant inhibitory effects on brook trout (Driscoll

et al., 1980) and white suckers (Baker and Schofield,

1982).

Mobilization of aluminum in Adirondack lakes (pH of 4.3

to 5) may be the primary factor limiting trout survival

there (Schofield and Trojnar, 1980). Freeman and Everhart (1971, 1973) studied the effects of aluminum hydroxide and aluminate (pH greater than 7.5) on rainbow

trout and discovered that the aluminate ion is acutely

toxic to trout at or above levels of 0.5 mg/L. In addition,

freshly precipitated aluminum hydroxide can cause

chronic injury in fish. Hall et al. (1985) conducted instream studies on the episodic effect of aluminum addition. They discovered that episodic additions of

aluminum can have significant biological, physical, and

chemical consequences in dilute, acidic surface waters.

An increase in aluminum under acidic conditions causes

the disruption of the ion regulatory mechanism and,

thus, the subsequent loss of tissue ions in various

aquatic biota. Species of fish (Buckler et al., 1987; Muniz

and Leivestad, 1980; Wood and McDonald, 1987),

daphnids (Havas, 1985), and immature aquatic insects

(Havas and Hutchinson, 1983; Witters et al., 1984) have

experienced net losses of sodium (Na+) and chloride

ions (Cl-

) under certain conditions in the presence of

aluminum. Ionic imbalances in fish resulting from aluminum exposure are usually accompanied by low pH and

alkalinity levels in surrounding waters (Muniz and

Leivestad, 1980).

Aluminum may also cause antagonistic effects to fish

when pH is less than 5 (Brown, 1983; Evans et al., 1988;

Havas, 1985; Hunn et al., 1987; Wood and McDonald,

1987). Baker and Schofield (1982) concluded that the

initial step toward aluminum toxicity involved the binding

of aluminum hydroxide to the gill surfaces.

Respiratory blockage in fish is another effect of increased aluminum concentrations. Fish may experience

symptoms such as elevated ventilation frequency, increased standard oxygen uptake rate, and a fall in dorsal aortic oxygen tension (Baker, 1982; Muniz and

Leivestad, 1980). This effect is normally observed at

circumneutral pH where aluminum exists as a colloidal

species (Malate, 1986; Ramamoorthy, 1988), or in the

presence of high calcium concentrations, where the effect on the ion regulatory mechanism is masked

(McDonald et al., 1983; Muniz and Leivestad, 1980;

Wood and McDonald, 1987).

74
Aluminum may exert indirect effects on aquatic organisms as well. The release of copper (Cu2+) from an

organic chelator due to competitive replacement by

aluminum was toxic to Scenedesmus (Rueter et al.,

1987). In addition, aluminum’s high affinity for phosphate and other compounds could remove nutrients

from the environment and thereby inhibit productivity

(Zarini et al., 1983; Sanville et al., 1976; Funk et al., 1977).

George et al. (1991) evaluated the toxicity of water

extracts from the discharged alum residuals of ten North

American WTPs. The receiving water samples from

each plant were adjusted to three pH levels for testing

purposes. Extracts from these samples were subjected

to a battery of toxicity tests, including the Selenastrum

capricornutum growth test, Ceriodaphnia dubia survival

and reproduction test, the fathead minnow survival and

growth test, a protozoan mortality test, and the Microtox

test. S. capricornutum was more sensitive than any

other test organism to the sludge extracts. Algal growth

inhibition was observed at pH 5 and generally not at

circumneutral pH. Dissolution of aluminum was observed in acidic and basic solutions; aluminum was only

slightly soluble in circumneutral solutions. Alum sludge

extracts were toxic to S. capricornutum at all receiving

water pH levels when the receiving water hardness was

less than 35 mg CaCO3/L.

5.1.2.2 Benthic Macroinvertebrates

Potential threats to benthic communities are magnified

when direct discharge occurs in quiescent bodies of

water. When discharging WTP residuals to a low-velocity languid stream or lake, mass balance equations (see

Section 5.2.1.3) do not apply. In this instance, a sludge

deposit made up of discharge effluent will accumulate in

the locus of the point of discharge. This poses a potential

threat to any benthic communities at, or in close proximity to, the discharge point. A major concern with direct

discharge is that residuals solids will cover the bottom

sediments, damaging the periphyton community. Sludge

deposits on sediments in receiving waters may limit

carbon sources on which macroinvertebrates feed, and

limit the available oxygen required for respiration. Areas

of lesser flow with more clay substrate provide a more

suitable benthic habitat than a sandy river channel does.

Historically, studies on the impacts of WTP effluent on

benthic communities in large- to mid-sized rivers have

found little evidence of related environmental degradation. A study by Evans et al. (1979) on the impact of

waste discharges on the Vermillion River from a water

works employing alum coagulation/filtration treatment

techniques, concluded that the influence of the waste

discharges on macroinvertebrates was imperceptible. A

later study by Evans et al. (1982) on waste discharges

from an alum coagulation/rapid sand filter plant on the

Mississippi River, found no marked environmental degradation to the river, as measured by sediment size

distribution and benthic macroinvertebrate abundance

and diversity.

A study on the influence of residuals discharge on

Crooked Creek in Illinois (Lin and Green, 1987) concluded that stream sediments immediately downstream

of the discharge outfall showed an increase in chemical

concentrations, a change in particle size distribution,

and a shift in the diversity and abundance of macroinvertebrates. The macroinvertebrate biotic index (MBI),

used by Illinois to indicate water quality on the basis of

the type of benthic life present, showed no change at

the sample stations immediately upstream and downstream of the water plant’s discharge (Illinois, 1987).

Studies on the effects of alum sludge on benthic communities provide additional information on aluminum

toxicity. Aluminum hydroxide has been shown to be toxic

to Tanytarsus dissimilis (a representative of the chirocomidae), at concentrations of 80, 240, and 480 mg/L

(Lamb and Bailey, 1981). A heavy floc layer on a lake

bottom could inhibit deposited eggs from reaching the

sediment and the surface.

An assessment of the effects of alum sludge discharges

into coastal streams (Roberts and Diaz, 1985) shows a

depression in the productivity of phytoplankton during

an alum discharge event. This was attributed to the high

level of suspended solids associated with the sludge

discharge. The researchers concluded that “this turbidity

effect in itself would argue for the cessation of sludge

discharge even in the absence of toxic effects” (Cornwell

et al., 1987). At plants where coagulation sludge is

allowed to accumulate in settling basins for several

months and is then discharged over short periods of

time, a substantial increase in total suspended solids

(TSS) and turbidity has been observed in the plant’s

receiving waters. Continuous withdrawal may minimize

this problem (Illinois, 1987).

5.1.2.3 Nonaluminum Substances

Residuals discharged from lime softening WTPs may

not adversely affect aquatic organisms due to the protection provided by high calcium and alkalinity levels.

Toxicity testing, however, of residuals extract obtained

at pH 8.3 from the Florence Water Treatment Plant (a

lime softening facility), Omaha, Nebraska, indicated

growth inhibition of S. capricornutum only in 50 and 100

percent extract solutions (George et al., 1995). Residuals extracts obtained at pH 6.0 also inhibited S. capricornutum growth in 12.5 percent extract solution plus

87.5 percent Missouri River water.

Most substances discharged into surface waters from

WTPs have established toxicity levels and are state

regulated. In-stream guidelines and standards are presented in Table 5-1. After characterizing the waste

75
Table 5-1. Possible In-Stream Water Quality Guidelines and Standards (Cornwell et al., 1987)

Guidelines Example Standards

Aquatic Life Chronic Criteria

Fresh mg/L Salt mg/L

Human

Healtha Stream Used for

Potable Water

Arsenic (dissolved) 72 63 2.2 ng/L 0.05 mg/L

Barium 1.0 mg/L

Beryllium 130 3.7 ng/L

Cadmium e1.16 (ln(hardness))-3.841 12 10.0 μg/L 0.01 mg/L

Chloride 250 mg/L

Chromium 7.2 54

 hexavalent, dissolved

 trivalent, active e0.819 (ln(hardness))+.537 170 μg/L

 total 0.05 mg/L

Copper 2.0 4(2y) 1.0 mg/L 

23(A)

Cyanide, free 4.2 0.57 20.0 μg/L

Fluoride 1.4 mg/L

Hydrogen sulfide 2.0 2.0

Iron

 total 1,000

 soluble 0.3 mg/L

Lead e1.34 (ln(hardness))-5.245 8.6 50 μg/L 0.05 mg/L

Manganese 

 total 100

 soluble 0.05 mg/L

Mercury 0.00057 0.1 146 ng/L 0.002 mg/L

Nickel (total) e0.76 (ln(hardness))+1.06 7.1 13.4 μg/L

Nitrate (as N) 10 mg/L 

Phenol 1.0 1.0 3,500 μg/L 0.001 mg/L

Selenium 35 54 10 μg/L 0.01 mg/L

Silver 0.01e1.72 (ln(hardness))-6.52 0.023 50 μg/L 0.05 mg/L

Sulfate 250 mg/L

TDS 500 mg/L

Zinc 47 58 5,000 μg/L 5.0 mg/L

Aldrin 0.03 0.003 0.074 ng/L

Chloride 0.0043 0.004 0.46 ng/L

Endrin 0.0023 0.0023 1.0 μg/L 0.0002 mg/L

Heptachlor 0.0038 0.0036 0.28 ng/L

Lindane 0.08 0.0016 0.004 mg/L

Methoxychlor 0.03 0.03 0.10 mg/L

Toxaphene 0.013 0.0007 0.71 ng/L 0.005 mg/L

DDT 0.001 0.001 0.024 ng/L 0.1 mg/L

Chloroform 1,240 0.19 μg/L

Radioactivity (Ra226+228) 5 pCi/L

Gross alpha particle activity (excluding

radon and uranium)

15 pCi/L

a Values given are the ambient water quality criteria for protection of human health for noncarcinogens, and for carcinogens the value is the

risk of one additional case of cancer in 1,000,000 persons.

76
stream quality and flow rate, a mass balance calculation

(see Section 5.2.1.3) should be conducted to determine

possible toxicity of certain substances.

5.1.2.4 Summary

Based on the findings of research on aluminum toxicity,

the following recommendations should be followed

when disposal to surface waters is being considered for

WTP residuals:

• Toxicity tests should be conducted on a variety of

representative aquatic organisms from the intended

receiving waters to fully determine potential toxic effects to the ecosystem.

• Alum sludges should not be discharged to acid

streams (pH less than 6); discharge into water with

circumneutral pH levels, however, should have minimal effects on primary production, and is acceptable.

• Caution should be taken in discharging alum sludges

to soft receiving waters (i.e., hardness less than 50

mg CaCO3/L).

• An environmental assessment should be undertaken

for any sludge disposal to surface waters. Factors

such as receiving water use, sediment structure,

water chemistry, system hydrology, and receiving

water biology need to be assessed to ascertain if

discharge of alum sludges into a specific surface

water will have a detrimental effect on the ecosystem.

5.2 Applications

5.2.1 Stream Hydraulics

The primary sources for residuals discharged to streams

typically include clarifier underflow sludge, filter backwash wastes, and sedimentation basin sludge. These

residuals are generated at varying rates. For example,

filter backwash wastes are generated at high flow rates

over short periods of time and may require equalization

before discharge; sedimentation basin sludge is discharged more frequently according to periodic basin

cleaning, unless the basins are equipped with mechanical sludge removal.

The impact that discharge of drinking water treatment

residuals has on a stream depends upon the frequency

of the discharge (ranging from daily to annually), the

variability of the stream flow, and the load discharged.

Under conditions of high velocity flow, more frequent

controlled discharge is possible because the residuals

will be suspended in the flow and dispersed downstream, diluting and minimizing the potential toxicity of

the residuals. As previously described, under conditions

of low velocity flow, high frequencies of discharge can

significantly affect the aquatic ecology of the stream

because the residuals will deposit in the vicinity of the

outfall. This type of deposit can be subject to scouring

during periods of high flow velocities.

5.2.1.1 Stream Characteristics and Flow Analysis

When considering the effects of a point source discharge on a stream, the principal stream characteristics

of interest include geometry (width and depth relations),

velocity, and flow. Stream flow, Q, is a function of the

average stream velocity, U, and cross-sectional area, A,

as:

Q = U(A) (Eq. 5-1)

Stream flow rates can be determined by several methods (Thomann and Mueller, 1987):

• Directly measuring cross-sectional area and average

flow velocity (as with a current meter).

• Tracking of markers or dyes over time.

• Obtaining stage-discharge relations from gauging

stations (operated by the U.S. Geological Survey along

most major rivers).

Given the concept of dilution, it is intuitive that the effects

of residuals discharges are most critical during low

stream flow periods. Consideration of minimum flow

involves an analysis of a stream’s flow frequency. Generally, most design work uses the minimum average

7-day flow expected to occur once every 10 years.

Minimum design flow is calculated using the following

equation:

minimum design flow = a Q b (Eq. 5-2)

where

a = number of days used in the average

Q = stream flow (see Eq. 5-1)

b = interval in years over which the flow is expected

to occur

The minimum design flow, therefore, is expressed as:

minimum design flow = 7(Q)(10) (Eq. 5-3)

Determinations of flow frequency can be made from

plotting a stream’s flow data and the percent of the time

that the stream will have a flow less than or equal to a

specific flow rate (Figure 5-1). From a flow frequency

analysis of the Schuylkill River in Philadelphia, Pennsylvania, for instance, the discharge of a periodic effluent

could be timed to coincide with the river’s high flow

periods to minimize effects of the discharge on the

receiving stream.

5.2.1.2 Length to Mixing

The concentration and flow of a pollutant into a stream

whose flow and other hydraulic properties are defined,

can be approximated (Thomann and Mueller, 1987).

This requires the assumption that water quality variables

77
are homogeneous laterally, across the stream, and vertically, with depth. In addition, it must be assumed that

the stream flow involves only advective flow—i.e., there

is no mixing in the longitudinal, or downstream, direction. True plug flow in streams is never attained, but

these assumptions are valid for a simple steady-state

analysis. Effluent discharge from a point source will

result in a plume that will gradually spread across the

stream and extend over the entire depth. The estimated

distance from point of discharge to the point of complete

mixing of effluent in the stream depends on whether the

discharge enters the stream from the side bank or from

midstream. For a side bank discharge, the length from

the source to the zone where the discharge is laterally

well mixed, Lm (in feet), is approximated by:

Lm = 2.6 U B2

/H (Eq. 5-4)

where

U = average stream velocity in feet per second

B = average stream width in feet

H = average stream depth in feet

For a midstream discharge, the mixing length is approximated by:

Lm = 1.3 U B2

/H (Eq. 5-5)

5.2.1.3 Flow and Mass Balance Calculations

Flow and mass balance calculations can be used to

estimate pollutant concentrations downstream from an

effluent discharge. To simplify mass balance calculations for point source effluent discharge to a stream,

assume that effluent mixing in the stream occurs instantaneously and that flow occurs only through advective

processes—i.e., there are no dead zones downstream

to cause flow dispersion, for example.

Figure 5-2 shows two scenarios for water supply and

effluent discharge sources for a WTP, where

Qu = upstream flow

Qi = influent plant flow

Qe = effluent plant flow

Qd = downstream flow

Cu = upstream pollutant concentration

Ce = effluent pollutant concentration

Cd = downstream pollutant concentration

The flow variables are in units of volume per time and

the concentration measurements are in units of mass

per volume. This results in units of mass per time, or

mass flux, in the mass balance calculations.

In the first scenario (common supply/discharge source)

depicted in Figure 5-2, the flow balance equation is:

Qd = Qu – Qi

 + Qe (Eq. 5-6)

The mass balance equation is:

QdCd = QuCu – Qi

Cu + QeCe (Eq. 5-7)

Typically, the upstream concentration of the pollutant is

omitted because it is much less than that of the effluent

or downstream concentration. The mass balance equation is then reduced to:

QdCd = QeCe (Eq. 5-8)

or

Cd = (QeCe)/Qd (Eq. 5-9)

Figure 5-2. Scenarios for mass balance calculations.

Figure 5-1. Flow frequency analysis of the Schuylkill River,

Philadelphia, PA; minimum 7-day average flow values, 1932-1964 (after Thomann and Mueller, 1987).

78
In the second scenario for water supply and effluent

discharge sources, the flow balance equation is:

Qd = Qu + Qe (Eq. 5-10)

The mass balance equation is:

QdCd = QuCu + QeCe (Eq. 5-11)

Again, assuming the upstream concentration is insignificant, the mass balance equation is reduced to the same

form as for the first scenario:

Cd = (QeCe)/Qd (Eq. 5-12)

If the effluent stream is continuous, the pollutant is conservative (its concentration does not degrade with time

or flow), and the river is assumed to flow as an ideal plug

flow reactor, then the downstream concentration, Cd, will

be constant for all downstream distances. Pollutant concentrations from WTP effluents, however, will decrease

downstream from the discharge point through advective

flow processes in the river, and chemical reactions,

bacterial degradations, or particulate settling of the pollutant in the water column. Typically, the concentration

decreases occur according to first order reaction kinetics. The concentration of a substance downstream is

then calculated as:

Cd = Cm[(-kX)/U] (Eq. 5-13)

where

Cd = downstream pollutant concentration

Cm = mixed effluent and upstream pollutant

concentration

k = first order reaction constant

X = distance downstream from the discharge point

U = river velocity

The change in pollutant concentration with downstream

flow is shown in Figure 5-3.

5.2.1.4 Water Quality Controls

Three factors influence the downstream fate of a pollutant: effluent flow rate, stream flow rate, and the decay

rate of the pollutant (Thomann and Mueller, 1987). By

modifying one of these variables, the downstream water

quality can be improved. Figure 5-4 shows the effects of

three techniques for controlling downstream water quality: decreasing effluent flow rate (a), making use of

increased upstream flow rate (b), and increasing decay

rate, k (c). With the first technique, if the effluent flow

rate is decreased and the effluent concentration remains

constant, the mixed pollutant concentration at the discharge point and the downstream concentration will be

reduced.

The second technique involves discharging during increased upstream flow rates, as from a storm event. In

this case, the discharged effluent will mix with a greater

flow of water in the stream and pollutant concentrations

near the discharge point will decrease. The disadvantage of this technique is that high stream flow rates will

carry pollutant concentrations farther downstream than

usual.

The third technique requires more sophisticated methods to modify the pollutant properties so that its decay

rate increases in the stream. The pollutant concentrations at the discharge point are unaffected by this technique, but the increased decay rate results in rapid

decreases in downstream pollutant concentrations.

5.2.2 Available Transport and Chemical

Models and Their Application

Numerous computer models are used to predict the

transport and fate of pollutants in natural systems. The

EPA Center for Exposure Assessment Modeling in Athens, Georgia, provides and supports public domain software that simulates air, water, and soil processes,

enabling the user to make risk-based decisions concerning environmental protection (Ambrose and Barnwell, 1989). The software is available to the user at no

cost, but considerable technical expertise is needed to

successfully use this software for problem-solving.

MINTEQ, for example, is one of six geochemicalequilibrium computer models routinely used to quantitatively evaluate the equilibrium behavior of inorganic and

organic constituents in various chemical environments

(U.S. EPA, 1991a; Bassett and Melchoir, 1990; U.S.

Figure 5-3. Decay of a nonconservative pollutant.

79
EPA, 1984a). MINTEQ describes the chemical conditions and reactions that control leaching. It then calculates the fate (or mass distribution) of dissolved and

solid phases of a pollutant under equilibrium conditions

in dilute aqueous systems. The model considers ion

speciation, adsorption, redox, gas-phase equilibria, and

precipitation and dissolution reactions in its calculations.

To obtain accurate information from the model, users

must be familiar with these processes and techniques

and be able to quantify them for a specific pollutant and

river system. There are currently no user-friendly aqueous chemical equilibrium models that are sophisticated

enough to make the necessary accurate downstream

water quality predictions.

5.3 Examples

An investigation of the adverse effects of alum sludge

on benthic invertebrates was conducted in 1991 by

Tennessee Technological University (TTU) for the

AWWA Research Foundation (George et al., 1991).

Three of the four following examples were summarized

from this research report. The fourth example was obtained from a report to the City of Phoenix, Arizona,

evaluating the discharge of three WTPs to the Salt River

Project canal system (HDR, 1995).

5.3.1 California Plant, Cincinnati Water

Works, Cincinnati, Ohio

5.3.1.1 Description

The California Plant in Cincinnati, Ohio, treats drinking

water at a rate of 618 m3

/min. The WTP consists of

primary coagulation-sedimentation, offstream storage

reservoirs, secondary coagulation-sedimentation, sand

filtration, and disinfection facilities (Figure 5-5). Primary

sedimentation is achieved using a 460 m3

/min facility

equipped with plate settlers to separate solids that are

continuously returned to the Ohio River. This process is

responsible for the bulk of the solids removal conducted

at the facility. The water from the storage ponds is

pumped to the secondary sedimentation facility, which

delivers a low turbidity water to the sand filters. The

storage reservoirs provide emergency capacity during

hazardous spill conditions that might occur in the Ohio

River system.

Aluminum sulfate and polymer are added to the raw

water. Ferric sulfate, powdered activated carbon (PAC),

and polymer are used intermittently. When needed, PAC

can be applied to the raw water; after the primary sedimentation basins and before pumping the water to the

storage reservoirs; and at the rapid-mix basin before the

secondary sedimentation basins. The plant can prechlorinate the raw water; add chlorine at the rapid-mix

basin between the reservoirs and the secondary sedimentation facility; chlorinate between the secondary

sedimentation basins and the sand filters; and

postchlorinate between the filters and the clearwell. Ferric sulfate, lime, and soda ash also can be added at the

rapid-mix basin prior to the secondary sedimentation

basins. The quantities of chemicals used at the California Plant from July 1988 through June 1989 are presented in Table 5-2.

Waste solids from the secondary sedimentation facility

and the backwash water are transported to the primary

sedimentation facility, which has plate settlers. An average of 5 m3

/min of the wasted sludge (0 to 4 percent

solids) from the primary sedimentation basins is discharged continuously to the Ohio River.

Data on the chemical composition of the Cincinnati

sludge and of the Ohio River at the point of discharge

are presented in Table 5-3. The Ohio River samples

tested by George et al. (1991) were moderately hard

(123 to 155 mg CaCO3/L); alkalinity ranged from 53 to

67 mg CaCO3/L; and total aluminum concentrations

Figure 5-4. Control techniques for improving downstream

water quality.

80
were less than 0.02 to 0.56 mg/L. The waste stream

from the WTP contained little aluminum (1.69 to 24.1

mg/L), and suspended solids varied from 170 to 4,272

mg/L. The Ohio River provided tremendous dilution of

the waste stream from the plant (2 x 10-5 m3

 sludge/m3

of river flow).

5.3.1.2 Toxicity

Samples of discharged sludge and receiving water were

taken for three different pH levels. Toxicity tests indicated that none of the sludge extracts produced a toxic

response in a protozoan (Tetrahymena) mortality test,

Microtox toxicity test, Ceriodaphnia toxicity test, or fathead minnow (Pimephales) toxicity test. At pH 5, however, the sludge extract was toxic to S. capricornutum

(Table 5-4) at a 50 percent and 100 percent extract

exposure. No adverse response was observed at pH 7.4

and 8.

5.3.1.3 Benthic Organisms

At the time it was studied, the California Plant discharged alum sludge to the Ohio River at a point just

above the confluence with the Little Miami River. PAC

was added as needed for taste and odor problems.

Table 5-2. Average Daily Chemical Use at California Plant, Cincinnati Water Works, Cincinnati, OH (George et al., 1991)

Month

Alum

(kg/d)

Polymer

(kg/d)

PAC

(kg/d)

Ferric Sulfate 

(kg/d)

Lime

(kg/d)

Fluoride

(kg/d)

Chlorine

(kg/d)

1988

July 6,882 0 970 0 6,729 652 2,970

August 6,313 0 6,628 877/15a 8,133 666 3,453

September 4,948 0 2,302/27a 766/13a 5,938 445 2,585

October 4,886 0 0 742/18a 4,695 425 1,876

November 4,895 207 0 646/5a 3,895 363 1,688

December 6,648 142 0 656/16a 4,501 398 1,546

1989

January 8,849 0 0 662/19a 4,751 405 1,426

February 7,470 986 0 965/13a 4,695 452 1,223

March 6,346 625 1,404 638/18a 4,171 429 1,110

April 6,864 405 3,103 0 4,156 465 1,132

May 7,555 574 1,986 618/16a 4,540 466 1,134

June 4,886 531 0 757/25a 4,702 554 1,806

Meanb 6,379 496 2,731 733 5,076 477 1,829

Figure 5-5. Flow schematic of California Plant, Cincinnati Water Works, Cincinnati, OH (George et al., 1991).

81
In January, March, and May 1989, water and sediment

samples were collected from the Ohio River to determine the impact of the sludge discharge on the benthic

communities. During January, aluminum concentrations

in the water column immediately above the sediments

ranged from 1.63 to 4.67 mg/L, with the highest concentration occurring at the point of discharge. Sediment

aluminum concentration (Figure 5-6) was also greatly

elevated at the point of discharge (23,560 mg/kg) and

gradually decreased to 1,882 mg/kg of dry sediment as

the sludge was dispersed downstream.

Aluminum concentrations in the water column immediately above the sediment collected in March ranged

from 3.92 to 8.96 mg/L. The two highest concentrations

of aluminum were found at 1,700 meters downstream

(8.69 mg/L) and 1,158 meters downstream (8.96 mg/L),

Table 5-4. S. Capricornutum Test Results on Alum Sludge

Extracts From California Plant, Cincinnati Water

Works, Cincinnati, OH (George et al., 1991)

Sample

Site pH

Percent 

Filtrate

Monomeric

Aluminum

Concentration

(mg M-Al/L)

Percent

Inhibition

Percent

Stimulation EC50

pH 5 12.5 <0.04 NE —a

25 0.05 NE —

50 0.10 35 —

100 0.20 70 —

pH 7.4 12.5 <0.04 NE —

25 <0.04 NE —

50 0.07 NE —

100 0.14 NE —

pH 8 12.5 0.05 NE —

25 0.10 NE —

50 0.19 50 —

100 0.38 NE —

a Insufficient data to compute EC50.

Key

NE = No statistically determined effect.

EC50 = Effective concentration producing a 50 percent reduction in

growth rate compared with the growth rate of control populations.

Figure 5-6. Sediment aluminum concentration (dry weight) from

Ohio River, Cincinnati, OH (George et al., 1991).

Table 5-3. Chemical Composition of Sludge and Ohio River Water Sampled at California Plant, Cincinnati Water Works,

Cincinnati, OH, September 21, 1988, December 18, 1988, and January 10, 1989 (George et al., 1991)

Sludge Ohio River

Parameter 9/21/88 12/18/88 1/10/89 9/21/88 12/18/88 1/10/89

Aluminum (mg/L) 24.1 23.6 1.69 <0.02 <0.02 0.56

Iron (mg/L) 12.0 13.4 0.84 0.16 <0.005 0.56

Manganese (mg/L) 3.59 <0.005 0.056 0.048 <0.005 0.22

Calcium (mg/L) 124.0 43 34 35 40 34

Magnesium (mg/L) 20.3 12.5 9.4 13.5 10.0 9.2

Total organic carbon (mg/L) 84 2.3 21 2.5 2.1 6.4

Alkalinity (mg CaCO3/L) 1,122 56 53 54 53 67

Total suspended solids (mg/L) 4,272 170 3,656 10.2 5.4 92

pH 7.8 7.0 7.4 7.6

82
and were approximately twice the aluminum concentrations measured at the upstream stations. Discharged

alum sludge floc may have been entrained in the water

currents. The greatest sediment concentrations of aluminum observed in March were collected near the point

of discharge. Concentrations of aluminum in the sediment ranged from 2,100 to 20,000 mg/kg dry weight.

Increased concentrations of aluminum at the point of

discharge (20,000 mg/L) were expected, due to settling

of suspended solids as the stream velocity decreases

toward the bank of the river. Aluminum concentrations

in the Ohio River during May remained fairly constant,

ranging from 0.94 to 1.27 mg/L. In May, aluminum concentrations in the sediment were highest near the shore

with less than 2,500 mg/kg measured in the deeper

channel.

The California Plant commonly added ferric sulfate,

along with increased amounts of aluminum sulfate, to

the raw water because of high turbidity following rain

events. The mean iron concentration in January was

1.79 mg/L (range 0.79 to 4.14 mg/L); it increased to 6.52

mg/L (range 2.55 to 7.90 mg/L) in March. By May (sampling period #3), the mean iron concentration dropped

to 1.92 mg/L (range 1.56 to 2.49 mg/L). Aluminum concentrations were also higher in March, increasing to a

mean of 5.69 mg/L (range 3.92 to 8.69 mg/L) from a

mean of 2.57 mg/L (range 1.63 to 4.67 mg/L) observed

on the previous sampling trip in January. The alum

sludge discharge, more dense with the addition of PAC,

apparently settled out quickly near the point of discharge

in January. Greater flow throughout the remainder of the

study period transported sediment deposits downstream

from the point of discharge.

Benthic macroinvertebrate data from the same sampling

location are presented in Table 5-5. Benthic samples

were not collected in May because of high flow velocities. Heavy scouring of the sandy bottom during high

flow was observed by Chisholm and Downs (1978), who

found damage to benthic communities resulting from the

relatively high flow velocities. All composited benthic

macroinvertebrate samples collected contained fewer

than 100 organisms. As the accuracy of the species

diversity indices is diminished when the samples contain

fewer than 100 organisms (U.S. EPA, 1983), no diversity

indices were calculated. Dipterans (Chironomidae) and

oligochaetes were the dominant taxa collected in these

samples.

No organisms were collected at the point of discharge

in January; this phenomenon was probably due to

sludge accumulations over the substrates. The temporary nature of these sludge deposits was evidenced by

the presence of a benthic community sampled at the

point of discharge in March. At the same time, no organisms were found in the sample taken 1,661 m downstream from the point of discharge. The lack of

organisms at this site immediately below the confluence

of the Little Miami and the Ohio Rivers was probably due

to high flow conditions in both water bodies and apparent scouring of the sandy bottom. Low numbers of taxa

and organisms at all stations indicate stress to the benthos in the study area. Lack of suitable substrate is

believed to be a major contributing factor to the low

species richness numbers of organisms in this reach of

the river.

In summary, the effects of alum sludge discharged from

the California Plant into the Ohio River could not be

separated from ambient impacts within the system. High

concentrations of sludge components in the study area

appear to be transitory in both the water column and

sediment. Benthic macroinvertebrate communities appear to exhibit transient impacts at the point of discharge

that were overshadowed by high-flow scouring from

storm events and lack of suitable habitat.

5.3.2 Ralph D. Bollman Water Treatment

Plant, Contra Costa Water District,

California

5.3.2.1 Description

The Ralph D. Bollman Water Treatment Plant, Contra

Costa Water District, California, treats water from the

Sacramento–San Joaquin River delta in northern California. Flow from the river delta is pumped into Mallard

Reservoir (capacity 13.8 million m3

), where it is held until

it is treated. The design capacity of the WTP is approximately 237 m3

/min, and the average production is 81

m3

/min. Water treatment consists of prechlorination;

alum coagulation; flocculation; sedimentation; chlorination prior to filtration; high-rate, dual-media filtration;

ammoniation; and pH control (Figure 5-7). At the time of

Table 5-5. Benthic Macroinvertebrates Collected From Site

CO on the Ohio River, Cincinnati, OH, 1989

(George et al., 1991)

Distance From Point of

Discharge

Number

of Taxa

Organism

Density (No.

Organisms/m2

)

Sampling Period 1, January 24-25, 1989

1,661 m downstream 3 113

1,158 m downstream 7 646

0 (point of discharge) —a —a

200 m upstream 4 1,227

Sampling Period 2, March 12-13, 1989

1,661 m downstream —a —a

1,158 m downstream 5 330

0 (point of discharge) 5 174

200 m upstream 7 322

Sampling Period 3, no sample

a Zero macroinvertebrates found.

83
this study, PAC was used to control taste and odor

problems. From November 1986 to December 1988, the

average dosage of PAC was 3.5 mg/L. The 1988 dosage

totaled approximately 146 metric tons. Also in 1988, an

average of 1,816 metric tons of alum and 182 to 227

metric tons of lime were added to the water.

Solids residuals from the sedimentation basins were

discharged daily to an embayment of Mallard Reservoir

for a 15-month period (October 1987 through December

1988). Filters were backwashed approximately once

every 50 hours and the water was wasted to the reservoir. During 1988, the backwash water was approximately 1 percent of the water treated. Approximately

1,504 metric tons of solids were discharged into Mallard

Reservoir. The water district did not waste solids from

the sedimentation basins to Mallard Reservoir after December 1988.

Sludge samples were collected from Mallard Reservoir

in October 1988 and February 1989 and contained from

803 to 1,450 mg/L aluminum and 196 to 319 mg/L iron.

Total suspended solids concentration in the sludge was

13,733 mg/L. Mallard Reservoir water contained 0.36 to

0.81 mg/L aluminum and was moderately hard (approximately 126 to 143 mg CaCO3/L). Chloride concentrations were high (159 to 217 mg/L) due to salt water

intrusion into the Sacramento–San Joaquin River delta.

5.3.2.2 Toxicity

Sludge extracts obtained with Mallard Reservoir water,

which was pH adjusted, were not toxic to bacteria (Microtox assay), the protozoan Tetrahymena ceriodaphnia, or fathead minnow. Sludge extracts obtained with

pH 5 reservoir water, however, were toxic to S. capricornutum (Table 5-6). An EC50 (the effective concentration

producing 50 percent reduction in growth rate compared

with the growth rate of control populations) of 22.5 percent was computed. Although the alum sludge extract

obtained at pH 7.5 stimulated algal growth at the lower

extract concentrations (i.e., 12.5 percent and 50 percent), no effect occurred at any concentration at pH 8.5.

5.3.2.3 Benthic Organisms

The Ralph D. Bollman WTP discharged wasted solids

daily into Mallard Reservoir from October 1987 through

December 1988. Taste and odor problems with the finished product water were controlled by the addition of

PAC, and the waste carbon was discharged to the

Table 5-6. S. Capricornutum Test Results on Alum Sludge

Extracts From Ralph D. Bollman Water Treatment

Plant, Contra Costa, CA (George et al., 1991)

Sample

Site pH

Percent

Filtrate

Monomeric

Aluminum

Concentration 

(mg M-Al/L)

Percent

Inhibition

Percent

Stimulation EC50

pH 5 12.5 <0.04 37 22.5%

25 0.05 54 —a

50 0.10 60 —a

100 0.20 100 —a

pH 7.5 12.5 <0.04 29 —a

25 <0.04 NE —a

50 0.05 43 —a

100 0.09 NE —a

pH 8.5 12.5 0.07 NE —a

25 0.14 NE —a

50 0.27 NE —a

100 0.54 NE —a

a Insufficient data to compute EC50.

Key

NE = No statistically determined effect.

EC50 = Effective concentration producing a 50 percent reduction in

growth rate compared with the growth rate of control populations.

Figure 5-7. Flow schematic of Ralph D. Bollman WTP, Contra Costa Water District, CA (George et al., 1991).

84
reservoir with the alum sludge. Sludge discharge to

Mallard Reservoir ended in December 1988.

Water and sediment samples were collected from Mallard Reservoir at three stations during February 1989.

Station 1 was located at the point of discharge, and

Stations 2 and 3 were located in the opposite corners of

the reservoir, outside the mixing zone. At the point of

discharge, aluminum concentration in the water column

above the sediments was 9.40 mg/L and averaged 0.36

mg/L within the bulk of the reservoir. The iron concentration was 2.80 mg/L at the point of discharge while an

average iron concentration of 0.15 mg/L existed in the

water column above the sediments in most of the impoundment.

Aluminum concentrations in the sediment (Figure 5-8)

were much higher at discharge point (102,000 mg/kg dry

weight) than at other locations within the reservoir (averaging 42,640 mg/kg). A thick layer of alum sludge

containing PAC appeared to be distributed over much of

the bottom surface of Mallard Reservoir.

Oligochaetes and amphipods dominated the benthic

macroinvertebrate communities at all stations. The point

of discharge had the highest number of taxa and also

the highest species diversity (Table 5-7). At the point of

discharge, eight taxa were identified and a Shannon

species diversity of 1.90 was computed. The other monitoring stations within the reservoir had five taxa with a

diversity of 1.16, and four taxa and a diversity of 0.62.

Species diversity and evenness values were lower at

Station 2 (500 organisms/m2

) and Station 3 (1,049 organisms/m2

), while the estimated number of organisms

per square meter, i.e., the density at these stations was

greater than at the sludge discharge location (113 organisms/m2

).

Impacted communities and stations usually exhibit a

reduced Shannon diversity value (Washington, 1984),

and values less than 3 indicate moderate to heavy

pollution (Goodnight, 1973). Margalef species richness

values were also highest at the discharge point (1.03)

and lowest at Station 3 (0.030). The major difference

between the benthic communities of the three sites was

the number of oligochaetes collected. Oligochaetes

composed 15 percent of the sample collected at Station

1, the point of discharge; 57 percent at Station 2; and

13 percent at Station 3. Fine particle, consolidated sediments apparently favored oligochaete assemblages at

the expense of aquatic insects (Pennak, 1989).

The low density of organisms at the point of discharge

(113 organisms/m2

) was indicative of moderate stress

on the benthic macroinvertebrate community and was

characteristic of a system subjected to toxic metal

wastes (Cairns et al., 1971; Young et al., 1975). Relatively high numbers of oligochaetes in the sample from

Station 2, along with a relatively low species diversity

and evenness value, also indicate a probable negative

impact on the ecological integrity of the benthic community (Washington, 1984; Cairns, 1977; Brinkhurst et al.,

1968). The Simpson heterogeneity index indicated

greatest species heterogeneity at the point of discharge

(0.84), where the highest number of taxa (eight) were

collected. Lowest heterogeneity (0.32) and lowest number of taxa (four) were found at Station 3.

In summary, data from the water and sediment field

samples taken at three stations at Mallard Reservoir

indicate moderate stress on benthic community structure, possibly resulting from alum sludge discharge and

habitat constraints. High concentrations of alum were

present in the water column and sediment of all stations,

with the highest concentration found at the point of

discharge. Assessment of benthic macroinvertebrate

Figure 5-8. Sediment aluminum concentration from Mallard

Reservoir, Concord, CA (George et al., 1991).

Table 5-7. Benthic Macroinvertebrates Collected From Mallard Reservoir at Site CC, Contra Costa Water District, Concord, CA,

February 21, 1989 (George et al., 1991)

Station

Number

of Taxa

Organism

Density (No.

Organisms/m2

)

Shannon

Diversity Index

Pielou

Evenness

Index

Margalef

Richness

Index

Simpson

Heterogeneity 

Index

1 8 113 1.90 0.91 1.03 0.84

2 5 500 1.16 0.72 0.45 0.63

3 4 1,049 0.62 0.45 0.30 0.32

85
data supports the observation of a stressed community

at all stations sampled.

5.3.3 Mobile Water Treatment Plant, Mobile,

Alabama

5.3.3.1 Description

The Mobile Water Treatment Plant has a capacity of 158

m3

/min and an average daily drinking water production

of 105 m3

/min. The raw water treated by the plant is

taken from the Big Creek Lake, which is a 1,457 hectare

(ha) reservoir. The water from the reservoir is very aggressive (Aggressive Index less than 10), with an average calcium concentration of only 2.9 mg CaCO3/L, a

bicarbonate concentration of 4 mg CaCO3/L, and a pH

of 6.2. Water is pumped from Big Creek Lake to a

ground-level storage basin from which water is pumped

to the plant. As water is transported from the storage

basin to the plant, it is disinfected by chlorine dioxide. At

the plant the water pH is raised by lime addition. Water

treatment includes coagulation by alum (average dose

24.8 mg/L), flocculation, sedimentation, rapid sand filtration and postchlorination (Figure 5-9). Zinc orthophosphate is added for corrosion control. The average mass

loading of aluminum to the system has been about 3.8

metric tons/day, and the average daily lime addition has

been 2.4 metric tons.

Settled solids from the sedimentation basins are wasted

daily for approximately 2.5 hours each morning at a rate

of 13 m3

/min. The filters are backwashed over a 3-hour

period every afternoon at a rate of approximately 2.1

m3

/min. The waste discharge goes into a small drainage

ditch that enters Three Mile Creek, which flows through

the city of Mobile, Alabama. At the time of the study,

Three Mile Creek water was soft (approximately 15 mg

CaCO3/L hardness) and had low alkalinity, ranging from

12 to 20 mg CaCO3/L. Aluminum in the Three Mile Creek

water samples ranged from less than 0.02 to 0.19 mg/L.

5.3.3.2 Toxicity

The protozoan mortality test using Tetrahymena indicated no toxic response of the organism to extracts from

the Mobile Water Treatment Plant waste sludge. Using

the Microtox assay, however, sludge extracts had a

15-minute EC50 equal to 85 percent of the filtered sludge

extract at pH 9. The predominant aluminum species at

pH 9 was most likely the aluminate ion. The Microtox

test showed no toxic effects to any of the remaining alum

sludge extracts. Mobile’s sludge extracts were significantly inhibitory to S. capricornutum at all concentrations of each pH level (Table 5-8).

5.3.3.3 Benthic Organisms

The brownish alum sludge discharge was clearly visible

for approximately 400 meters downstream of the point

of entry before mixing with the receiving water. During

water and sediment monitoring in January 1989, sulfate

concentrations in Three Mile Creek ranged from 2.7

mg/L upstream from the point of discharge to 18.7 mg/L

at the point of discharge. TSS also followed this trend in

January when concentrations ranged from 3.2 mg/L

upstream (Station 4) to 185 mg/L at point of discharge

(Station 3). Likewise, aluminum concentrations ranged

from 1.21 mg/L at Station 4 to 5.26 mg/L at Station 3 in

January.

While the aluminum concentration upstream from the

point of discharge ranged from 1.21 (January) to 0.02

Figure 5-9. Flow schematic of Mobile WTP, Mobile, AL (George et al., 1991).

86
mg/L (February), the downstream mean aluminum concentration in January was 12.6 mg/L, decreasing to 0.05

mg/L in February, and increasing to 1.09 mg/L in March.

Downstream mean TSS concentrations were also highest in January 78.2 mg/L, dropping to 5.2 mg/L in February. Iron exhibited a similar trend, as downstream

mean concentration was 2.59 mg/L in January, 0.30

mg/L in February, and 0.96 mg/L in March.

The data gathered suggest the possibility that deposition of aluminum from the plant’s alum sludge occurred

in sediment downstream from the point of discharge

(Figure 5-10). Aluminum concentrations in the sediment

at the downstream Station 2 (2,208 mg/kg) were higher

than the concentrations found at the point of discharge

(1,236 mg/kg).

Benthic macroinvertebrate data are presented in Table

5-9. Oligochaetes and dipterans (Chironomidae) were

the dominant taxa collected in Three Mile Creek, in

addition to trichopterans (Cheumatopsyche sp.). The

upstream monitoring station had the greatest number of

taxa and the greatest number of organisms for two of

the three sampling events. In February, the density of

organisms ranged from 4 organisms/m2

 to 3,606 organisms/m2

 at Station 4. By March, the number of organisms per square meter ranged from 41 at the point of

discharge to 3,641 at the upstream reference station. In

January, the lowest number of organisms was recorded

at the point of discharge. The alum sludge discharged

by the Mobile Water Treatment Plant apparently had an

adverse effect on the benthic macroinvertebrate density

of Three Mile Creek at and below the point of discharge.

A study by Lin (1989) also found an impact to benthic

macroinvertebrates from alum sludge discharge.

In summary, the alum sludge discharged by the Mobile

Water Treatment Plant appeared to have an adverse

effect on the benthic macroinvertebrate community of

the receiving stream at and below the point of discharge.

Moderate recovery of the macroinvertebrate community

appeared to have occurred approximately 1,200 meters

below the point of discharge.

Figure 5-10. Sediment aluminum concentration from Three Mile

Creek, Mobile, AL (George et al., 1991).

Table 5-9. Benthic Macroinvertebrates Collected From Three

Mile Creek, Mobile, AL, 1989 (George et al., 1991)

Station

Number

of Taxa

Organism

Density (no.

organisms/m2

)

Sampling Period #1 - January 19, 1989

1 - ~3 km downstream location

from point of discharge

6 1,570

2 - ~0.5 km downstream from

point of discharge

6 54

3 - point of discharge 6 14

4 - upstream monitoring station 11 517

Sampling Period #2 - February 14, 1989

1 9 529

2 8 57

3 22 4

4 21 3,606

Sampling Period #3 - March 21, 1989

1 12 335

2 3 167

3 5 41

4 15 3,641

Note: Samples were collected using a Surber sampler and consisted

of five replicates per station. The numbers of taxa and organisms are

the summations of the numbers found in each replicate sample for

each site.

Table 5-8. S. Capricornutum Test Results on Alum Sludge

Extracts From the Mobile, AL, Water Treatment

Plant (George et al., 1991)

Sample

Site pH

Percent

Filtrate

Monomeric

Aluminum

Concentration 

(mg M-Al/L)

Percent

Inhibition EC50

pH 6.7 12.5 0.04 18 37%

25 0.08 27 —a

50 0.17 62 —

100 0.34 92 —

pH 8 12.5 0.04 30 19%

25 0.08 63 —

50 0.17 95 —

100 0.34 99 —

pH 9 12.5 1.34 99.6 —

25 2.69 100 —

50 5.38 100 —

100 10.76 100 —

a Insufficient data to compute EC50.

Key

EC50 = Effective concentration producing a 50 percent reduction in

growth rate compared with the growth rate of control populations.

87
5.3.4 City of Phoenix Utility, Phoenix, Arizona

5.3.4.1 Description

The Val Vista, Deer Valley, and Squaw Peak WTPs,

which are owned and operated by the City of Phoenix,

Arizona, directly discharge their waste residual solids

into the Salt River Project (SRP) canal system (Figure

5-11). The SRP also serves as the raw water source for

these WTPs. EPA has determined that discharges to the

SRP canals, considered waters of the United States, are

subject to NPDES permitting requirements (HDR, 1995).

Water in the SRP is a combination of the Salt and Verde

Rivers. Furthermore, the SRP canal system is interconnected with water from the Central Arizona Project

(CAP). The quality of raw water treated by the WTP

varies according to the proportion of water received from

each of the aforementioned sources. Table 5-10 presents the average raw water quality for each of these

three sources.

Table 5-11 shows the flow rates and average turbidity

data for each facility. Even though Deer Valley and

Squaw Peak WTPs obtain raw water from the same

source (the Arizona Canal), the average raw water turbidities at the Deer Valley WTP were higher than turbidity levels treated by the Squaw Peak plant. The higher

raw water turbidity at the Deer Valley WTP has been

attributed to the residuals solids discharged into the

canal by the Squaw Peak WTP.

The three WTPs each use conventional treatment, employing grit removal, presedimentation, chemical addition, rapid mixing, flocculation, final sedimentation,

filtration, and disinfection. Direct discharge of residuals

solids into SRP was an acceptable practice at the time

the facilities were constructed and, therefore, no solids

handling facilities were constructed at any of the WTPs.

The following residuals solids were in waste streams

discharged to the canal system from the following water

treatment processes:

• Raw water solids (usually without chemical addition)

from grit and presedimentation basins.

• Alum residual solids from final sedimentation basins.

• Wasted washwater from filter backwashing and filterto-waste flows (i.e., filtered water immediately after a

filter is backwashed when turbidities exceed drinking

water standards).

• Miscellaneous waste streams from leaking valves,

basin and reservoir overflows, surface drainage, and

sample sink return flows (no chemical addition except

for chlorine residuals).

Currently, residuals solids from the three WTPs are

being discharged directly into the canals without equaliFigure 5-11. Location of City of Phoenix WTPs, Phoenix, AZ (HDR, 1995).

88
zation or treatment. Characteristics of the waste

streams from each of the WTPs are presented in Tables

5-12, 5-13, and 5-14. The solids mass loading discharged from each treatment process was computed

based on the flows and total suspended solids data

(Table 5-15). The Phoenix Water Service Department

funded a study to determine the effects of these residuals solids on the hydraulics, water quality, and biological

communities of the SRP canal system (HDR, 1995).

5.3.4.2 Effects of Discharged Residuals Solids

on Canal Hydraulics and Sediment Load

Solids discharged by the Val Vista WTP to the South

Canal are permitted through an agreement between the

City and SRP. The City pays SRP’s costs for dredging

sediment from a specific length of canal downstream

from each discharge point. These portions of the canal

are drained and dried annually. During this time, the

sediments are removed and placed on the canal banks

to dry. In 1995, the annual cost to the City ranged from

$15,000 to $50,000. If the discharged solids settle and

are retained in the canal, the resulting sediments may

affect the canal hydraulics.

The U.S. Army Corps of Engineers’ HEC-2 computer

models of canal physical and hydraulic characteristics

were used to determine the impact of residuals solids

on the SRP canal system. Computer simulations were

compared with information from the SRP on water level

profiles corresponding to specific flow rates in the canal.

The SRP also included water surface elevations corresponding to design capacities within the Arizona Canal

(19.8 m3

/sec at Station 38+862 to Station 73+963) and

South Canal (41.1 m3

/sec at Station 0+000 to Station

25+700). The velocities in the canal system varied

slightly but were generally about 94 cm/sec. The HEC-2

hydraulic analysis compared favorably with independently measured flow and cross-section data.

HDR (1995) evaluated the settling characteristics of the

discharged solids to determine if they would settle and

be retained in the canal system. According to Table 5-16,

81.4 percent of the solids measured were finer than a

#200 sieve. Using the U.S. Army Corps of Engineers’

HEC-6 computer program, transport of solids in the

canals was assessed. HEC-6 uses the same input

scheme as HEC-2 to describe the hydraulic and geometric characteristics of the canals. In addition, grain

size distribution data (Table 5-16) and total volume of

solids entering the canal are entered into the program.

The model calculates the volume of each grain size that

will settle out at each canal cross section. The canals

being evaluated are divided into reaches and the model

predicts the solids mass load for each of the reaches

(Table 5-17).

Computer results indicate that the material finer than a

#200 sieve (approximately 81 percent of the particles)

would be transported through the canal system. These

results were based on the assumption of maintaining a

water velocity of 58 cm/sec or greater. Localized areas

exist—eddies caused by intakes, delivery and dispersion structures, control structures, bridge piers, and

Table 5-10. Typical Canal Source Water Characteristics 

(HDR, 1995)

Parameter

(mg/L unless

noted otherwise)

Verde River

Below Bartlett

Dam, SRP

Canala

Salt River

Below Stewart

Mountain

Dam, SRP

Canala

Colorado

River Below

Parker Dam,

CAP Canala

Arsenic 0.013 0.004 0.003

Barium 0.1 0.1 0.7b

Bicarbonate 235 160 155

Boron 0.19 0.19 0.19b

Cadmium 0.006 0.007 0.001

Calcium 43 50 82

Carbonate 2 0 0

Chloride 19 235 91

Chromium 0.004 0.002 0.006b

Copper 0.009 0.009 0.006b

Fecal coliform

(cfu/mL)f 13 7 4

Fluoride 0.3 0.4 0.4 

Hardness 

(as CaCO3)

212 180 330

Hardness 19 49 203

Iron 0.19 0.19 0.13

Lead 0.07 0.07 0.01

Magnesium 26 14 30

Manganese 0.09 0.06 0.02

Mercury 0.00026 0.00004 0.00003

Nitrate (as N) 0.4 0.1 0.1c

pH (units) 8.0 7.7 8.0

Phosphorus 0.21 0.22 0.02

Potassium 3.4 5.8 5.1

Selenium 0.00006 0.00003 0.00003

Specific

conductance

472 1,138 1,083

Silver 0.00001 0.0017 0.0001b

Sodium 30 161 104

Sulfate 53 51 294

Total dissolved

solids

280 5,007 721

Total organic

carbon

3.5d 3.5d 5.2e

Turbidity (NTU)g 58 11 4c

Zinc 0.04 0.033 0.02

a National Stream Quality Accounting Network (NASQAN) Period of

Record, March 1975 to September 1979. Information on this water

source is provided because CAP water can be released to the Salt

River and be carried in the SRP canal. b Period of Record, Water Year 1979 only.

c

 Combined nitrate-nitrite.

d Water Quality Master Plan.

e Tucson Water, Period of Record: October 1986 to August 1987.

f

 cfu/mL: colony forming units per milliliter.

g Turbidities can range from as low as 1 NTU to as high as 500+

NTUs.

89
Table 5-11. Summary of Plant Flows and Turbidity Data (HDR, 1995)

Average Maximum Peak

WTP

Flow

(m3

/min)

Turbidity

(NTU)

Flow

(m3/min)

Turbidity

(NTU)

Flow

(m3/min)

Turbidity

(NTU)

Val Vista 315 11 368 20 368 40

Squaw Peak 184 10 368 20 368 40

Deer Valley 210 20 315 40 315 40

Table 5-12. Val Vista WTP Discharge Stream Characteristics (HDR, 1995)

Discharge Stream

Parameter Raw Water

Grit Basin

Discharge

Presed.

Basin

Blowdown

Final Sed.

Basin

Blowdown

Filter Waste

Washwater

Leakage and

Drains

Operational frequency Continuous Intermittent 1 hr/day 1–4 hr/day 4 times/day Continuous

Flow (m3

/min) 315 0.68 0.63 3.4 7.9 5.0

Turibidity (NTU) 11 N/A N/A N/A N/A N/A

Total suspended solids (mg/L) 20 16 1,600 1,400 55 3 

Total aluminum (mg/L) <0.05 N/A 19 1,800 165 0.010

Note:

1. To be reduced under current plant modifications.

2. Storm drainage not included.

Table 5-13. Squaw Peak WTP Discharge Stream Characteristics (HDR, 1995)

Discharge Stream

Parameter Raw Water

Presed. Basin

Blowdown

Final Sed. Basin

Blowdown

Filter Waste

Washwater Leakage

Operational frequency Continuous 3 times/day 0.5 hr/day 3 times/day Continuous

Flow (m3

/min) 184 1.6 1.3 1.3 1.3

Turbidity (NTU) 10 N/A N/A N/A N/A

Total suspended solids (mg/L) 30 130 3,900 360 N/A

pH 7.8 N/A N/A N/A N/A

Dissolved aluminum (mg/L) 0.2 0.2 0.2 0.6 N/A

Table 5-14. Deer Valley WTP Discharge Stream Characteristics (HDR, 1995)

Discharge Stream

Parameter Raw Water

Grit Basin

Discharge

Presed. Basin

Blowdown

Final Sed. Basin

Blowdown

Filter Waste

Washwater Leakage

Operational frequency Continuous 3 times/year 3 times/week 2 times/week 6 times/day Continuous

Flow 210 m3

/min 757 m3/

discharge

2,200 m3/

discharge

1,500 m3/

discharge

2.4 m3/min 5.3 m3/min

Turbidity (NTU) 20 >1,000 >1,000 >1,000 71.25 4.4

Total suspended solids (mg/L) 39 4,900 21,000 9,400 264 14.5

Settleable solids (mg/L) 0.16 18 290 310 35 0.27

pH 8.4 7.65 7.83 7.45 N/A

Temperature (°C) 24 24 26 28 N/A

Total dissolved solids (mg/L) 732 760 801 813 731 740

Total aluminum (mg/L) 1.27 240 769 704 45.3 0.91

90
mossy areas—where the stream velocity may fall below

58 cm/sec and increased deposition might occur.

The City of Phoenix investigated the concentration of

suspended solids in the Southern Canal following discharge of waste residuals solids from the Val Vista WTP.

The sample collected upstream from the point of discharge was a grab sample, whereas the remaining samples were time-weighted composite samples. The City

discovered that the TSS concentration at the point of

discharge in the canal ranged from 200 mg/L (presedimentation blowdown) to 1,600 mg/L (final sedimentation

blowdown). At 25 m downstream from the discharge

point, 50 percent of the solids had been removed, and

at 100 m downstream at least 70 percent of the solids

had been removed from the water column. Analysis of

suspended solids in water samples collected 500 m

downstream indicated a TSS reduction of 78 to 98

Table 5-15. Existing Discharge Quantities (HDR, 1995)

Val Vista WTP Squaw Peak WTP Deer Valley WTP

Discharge

Stream

Plant

Flow

Cond

 Flow

 (m3

/min) Dur.

(min)

Solids

(kg/day)

Flow

(m3

/min)

Dur.

(min)

Solids

(kg/day)

Flow

(m3/min)

Dur.

(min)

Solids

(kg/day)

Grit basin drain Ave. 0.68 Cont. 27

Max. 0.68 Cont. 50 No grit basin drain Intermittent dischargea

Peak 0.68 Cont. 100

Presediment

basin blowdown

Ave. 0.68 120 24,500 1.6 60 181 0.37 19 13,000

Max. 0.63 120 4,500 1.6 60 726 0.37 20 44,000

Peak 7.6 120 9,100 1.6 60 1,450 0.37 30 44,000

Final sediment

basin blowdown

Ave. 3.4 120 12,700 1.3 60 10,400 0.47 31 3,900

Max. 3.4 120 24,500 1.3 60 31,000 0.68 35 14,000

Peak 6.6 120 48,100 2.6 60 62,000 0.68 35 14,000

Filter waste

washwater

Ave. 12.6 8 1,700 1.3 10b 1,000 2.4 20b 900

Max. 12.6 30b 3,300 5.3 15b 4,400 5.8 20b 3,200

Peak 19.7 30b 6,400 10.0 15b 8,600 5.8 20b 3,200

Leakage and

drains

Ave. 0.18 Cont. 1.4 Included in final sediment

basin blowdown

5.3 cont. 15

Max. 0.18 Cont. 2.3 5.3 Cont. 15

Peak 0.18 Cont. 5.4 5.3 Cont. 15

Totals Ave. 17.6 17,200 4.2 12,000 8.5 18,000

Max. 17.6 32,200 8.1 36,000 12.1 61,000

Peak 33.9 63,500 14.2 73,000 12.1 61,000

a Grit basin discharges are estimated to be 760 m3 per event with a solids content of 3,600 kg. This discharge will be eliminated in upcoming

modifications. b Data obtained through telephone conversations with plant managers.

Table 5-16. Grain Size Distribution for WTP Residuals 

(HDR, 1995)

Sieve Size Retained (%) Passing (%)

#4 0 100 

#8 2 98

#10 1 97

#16 4 93

#30 5 88

#40 2 86

#50 1 85

#100 2 83

#200 1 82

Pass #200 81.4

Table 5-17. Estimated Amount of Solids Deposited in the

Canal System, Phoenix, AZ (HDR, 1995)

Suspended Solids

Load

WTP

Entering

Reach

(mt/day)

Existing

Reach

(mt/day)

Solids

Deposited

in Reach

(mt/year)

Length of

Canal

Reach (m)

Squaw Peak 24.5 22.4 762 10,700

Deer Valley 17.7 17.3 132 3,500

Val Vista 17.0 12.2 73 7,800

91
percent and TSS levels 3 to 45 times the concentration

measured upstream of the point of discharge.

5.3.4.3 Impact on Aquatic Organisms

The Arizona Game and Fish Department electrofished

the Arizona Canal between October 1992 and April

1993. The results of the electrofishing are presented in

Table 5-18. Ctenopharyngodon idellus (white amur or

grass carp) has been used for the control of aquatic

vegetation and, therefore, is probably the most economically important species.

Indigenous planktonic invertebrates were assumed to

have originated from the Salt River reservoir system.

Analysis of zooplankton in Canyon Lake by McNatt

(1977) revealed that at least 15 cladoceran, 5 copepod,

and 1 rotifer species may inhabit the rivers and canals.

Among the cladocera identified were Bosmina longirostris, Daphnia parvula, Daphnia pulex, Daphnia galeata,

Daphnia ambigua, Diaphanosoma leuchtenbergianum,

Ceriodaphnia lacustris, Chydorus sphaericus, and

Ilyocryptus sp. Copepod species found include Paracyclops fimbriatus, Acanthocyclops vernalis, Diacyclops

bicuspidatus, and Diaptomis siciliodes. Asplanchna priodonta was the identified rotifier. In the Salt River system benthic communities, several species have been

identified: Chironomus, Cryptochironomus, Polypedium,

Procladius, Tanypus, two tubificid worms (Branchiura

and Limnodrilus), the Asiatic clam (Corbicula malinensis), a crayfish (Oroconectes virilus) and two sponge

species. HDR (1995) did not conduct a study of the

phytoplankton present in the canal system. Based on

another analysis of water in the SRP reservoirs (Olsen,

1975), however, HDR inferred that a reasonable representation of flora in the canals would include Cyclotella,

Stephanodiscus, Synedra, Spirulina, Anabaenopsis, Dinobryon, Glenodinium, Phacotus, Closteriopsis, Tetrastrum, Tetraedron, and Ankistrodesmus.

The primary pollutants found in the WTP wastestreams

that may adversely affect aquatic organisms were arsenic, barium, cadmium, chromium, mercury, selenium,

and trihalomethanes. Barium is only toxic to freshwater

aquatic life at levels greater than 50 mg/L (U.S. EPA,

1986c) and was not viewed as a problem by the consultants. Concentrations of the remaining pollutants in the

discharged waste streams are presented in Table 5-19.

An evaluation of the pollutants in the City of Phoenix’s

waste streams revealed that 85 to 98 percent of the

metals discharged were in particulate form and would

ultimately settle once the turbulence was reduced downstream (HDR, 1995). HDR (1995) used a simple mathematical model to estimate the concentration of pollutants

in the water column downstream from the discharge

points. The assumptions were: 1) complete mixing of the

pollutants occurred in the volume of water passing the

Table 5-18. Total Number of Fish Caught by Electrofishing, Arizona Canal, Phoenix, AZ, October 15, 1992, to April 30, 1993 

(HDR, 1995)

Common Site #1 Site #2 Site #3 Site #5 Site #7 Total

Bluegill 10 0 0 0 0 10

Carp 1 1 1 1 1 5

Channel 5 35 1 2 20 63

Desert sucker 35 98 474 300 91 998

Goldfish 100001

Largemouth 20 7 3 7 0 37

Mosquitofish 0 0 0 0 7 7

Rainbow trout 010102

Red shiner 225 223 48 40 33 599

Roundtail 0 2 15 1 5 23

Smallmouth 100001

Sonora sucker 82 157 467 571 518 1,795

Threadfin 579 97 29 62 33 800

White amur 97 4 15 12 5 133

Yellow bass 11 4 0 3 2 20

Yellow 2 14 3 0 0 19

Site 1: From 67th Ave. downstream to the Skunk Creek drain gates.

Site 2: From the corner of 43rd Ave. and Peoria, downstream to the corner of 51st Ave. and Cactus.

Site 3: From the 19th Ave. bridge downstream to the Black Canyon freeway bridge, which lies between Peoria and Dunlap.

Site 5: From the 68th Street bridge downstream to the water control structure located near the 56th Street bridge.

Site 7: From Pima Road bridge downstream to Hayden Road, which lies between Indian Bend Road and McDonald Drive.

92
discharge point by the end of a 24-hour period; and 2)

no volume reduction was made to compensate for sedimentation. Using data in Table 5-19 to compute daily

pollutant mass loadings, average and maximum daily

pollutant contributions to the receiving stream were

computed for typical flows in the Arizona Canal (213

m3

/sec) and Southern Canal (442 m3/sec). HDR (1995)

also computed the pollutant concentration at a low flow

condition of 25 percent of the typical flow. The results

of these computations are presented in Tables 5-20

and 5-21.

Under maximum mass discharge conditions and typical

canal flow, the computed aluminum concentration in the

receiving water ranged from 1 mg/L (Val Vista WTP) to

45 mg/L (Deer Valley WTP). Estimated aluminum concentration during low flow would be 179 mg/L below the

Deer Valley point of discharge. Deer Valley’s residuals

solids would produce the highest concentration of metals in the canal water column (see Tables 5-20 and

5-21). The estimated increase in arsenic, cadmium,

chromium, and selenium in the water downstream from

Squaw Peak WTP would be less than 1 μg/L.

Metal Toxicity

Environmental factors may affect aluminum toxicity to

aquatic organisms. High calcium concentrations protect

against the effect of aluminum on the ion regulatory

mechanism of rainbow trout (McDonald et al., 1983;

Muniz and Leivestad, 1980), brown trout (Wood and

McDonald, 1987) and Daphnia magna (Havas, 1985).

Furthermore, natural organic compounds can complex

with aluminum and protect aquatic organisms (Birchall

et al., 1989; Driscoll et al., 1980; Karlson-Norrgren et al.,

1986). Increases in hardness and alkalinity reduce the

Table 5-19. Estimated WTP Discharge Stream Pollutant Concentrations (HDR, 1995)

WTP

Total

Discharge

(m3

/min)

TSS

(mg/L)

Al

(mg/L)

As

(mg/L)

Cd

(mg/L)

Cr

(mg/L)

Hg

(mg/L)

Se

(mg/L)

THM

(mg/L)

Squaw Peak

Presedimentation

Average 0.05 8,400 0.2 0.015 0.005 0.01 0.0002 0.005 0.100

Maximum 0.18 9,400 0.2 0.015 0.005 0.01 0.0002 0.005 0.100

Final sedimentation

Average 2.79 5,000 840 0.005 0.005 0.01 0.0002 0.005 0.136

Maximum 10.75 5,000 406 0.005 0.005 0.01 0.0002 0.005 0.136

Filter back wash

Average 4.73 190 8 0.010 0.005 0.01 0.0002 0.005 0.100

Maximum 18.92 280 11 0.010 0.005 0.01 0.0002 0.005 0.100

Deer Valley

Presedimentation

Average 0.03 14,400 0.2 0.428 0.036 0.30 0.0045 0.005 0.026

Maximum 0.16 9,400 0.2 0.428 0.036 0.30 0.0045 0.005 0.026

Final sedimentation

Average 2.37 4,700 560 0.478 0.028 0.34 0.0012 0.029 0.126

Maximum 9.12 4,700 560 0.478 0.028 0.34 0.0012 0.029 0.126

Filter back wash

Average 4.21 170 14 0.056 0.005 0.01 0.0002 0.005 0.061

Maximum 15.77 280 14 0.056 0.005 0.01 0.0002 0.005 0.061

Val Vista

Presedimentation

Average 0.05 7,800 0.2 0.005 0.005 0.01 0.0002 0.005 0.100

Maximum 0.16 7,800 0.2 0.005 0.005 0.01 0.0002 0.005 0.100

Final sedimentation

Average 2.68 4,800 214 1.170 0.005 0.07 0.0002 0.005 0.220

Maximum 9.59 5,000 300 1.600 0.005 0.10 0.0002 0.005 0.300

Filter back wash

Average 8.67 100 5 0.005 0.005 0.01 0.0002 0.005 0.100

Maximum 23.13 150 16 0.005 0.005 0.01 0.0002 0.005 0.100

93
degree of toxicity of alum sludge in water (George et al.,

1991). HDR (1995) concluded that with the neutral pH

levels, high alkalinity (130 to 185 mg CaCO3/L) and

hardness (180 to 212 mg CaCO3/L) of the receiving

water, aluminum toxicity to aquatic organism in canal

water should be minimized or eliminated.

Table 5-22 presents acute arsenic toxicity data for several

aquatic organisms. Estimated arsenic levels in the canal

water column downstream from a WTP discharge were

less than 0.2 mg/L. Arsenic does not appear to bioaccumulate in fish (U.S. EPA, 1986c). Based on the low bioaccumulation potential, short half-life of arsenic, and the

toxicity data, HDR (1995) concluded that arsenic toxicity

should not be a problem with aquatic organisms.

Cadmium can bioaccumulate and cause damage to fish

vital organs (Kay, 1986). Acute exposure to cadmium

can result in respiratory failure caused by damage to the

gills. Data on toxicity of cadmium to aquatic organisms

are presented in Table 5-23. As water hardness, alkalinity, salinity, temperature, and organic matter increase,

the toxicity of cadmium to aquatic organisms decreases

(Sorensen, 1991).

Hardness has a similar effect on chromium toxicity to

aquatic organisms (Mance, 1987; U.S. EPA, 1986c).

Table 5-20. Daily Resultant Pollutant Concentration Contribution to Canal at Typical Flow, Phoenix, AZ (HDR, 1995)

Resultant Concentration Contribution (mg/L)

Squaw Peak

Canal Flow

(m3

/s)

TSS

(mg/L)

Al

(mg/L) As Cd Cr Hg Se THM

Average 213 13 2 0.05 0.03 0.06 0.001 0.03 0.75

Maximum 213 51 4 0.20 0.12 0.25 0.005 0.12 2.83

Resultant Concentration Contribution (mg/L)

Deer Valley

Canal Flow

(m3

/s)

TSS

(mg/L)

Al

(mg/L) As Cd Cr Hg Se THM

Average 213 10 1 1.20 0.1 7.2 0.03 0.8 4.7

Maximum 213 41 45 46 2.9 28 0.13 2.9 18 

Resultant Concentration Contribution (mg/L)

Val Vista

Canal Flow

(m3

/s)

TSS

(mg/L)

Al

(mg/L) As Cd Cr Hg Se THM

Average 442 6 0 2 0.0 0.1 0.001 0.0 0.6

Maximum 442 21 1 6.3 0.1 0.4 0.003 0.1 2.1

Table 5-21. Daily Resultant Pollutant Concentration Contribution to Canal at 25% of Typical Flow, Phoenix, AZ (HDR, 1995)

Resultant Concentration Contribution (mg/L)

Squaw Peak

Canal Flow

(m3

/s)

TSS

(mg/L)

Al

(mg/L) As Cd Cr Hg Se THM

Average 53 51 8 0.20 0.12 0.26 0.005 0.12 3

Maximum 53 204 16 0.80 0.50 1.00 0.020 0.50 11

Resultant Concentration Contribution (mg/L)

Deer Valley

Canal Flow

(m3

/s)

TSS

(mg/L)

Al

(mg/L) As Cd Cr Hg Se THM

Average 53 41 5.0 4.6 0.3 29 0.11 3.1 18

Maximum 53 164 179 186 12 111 0.5 12 72

Resultant Concentration Contribution (mg/L)

Val Vista

Canal Flow

(m3

/s)

TSS

(mg/L)

Al

(mg/L) As Cd Cr Hg Se THM

Average 110 23 1.0 7.1 0.09 0.38 0.004 0.09 2.4

Maximum 110 86 5.3 25.3 0.27 1.76 0.011 0.26 8.5

94
Chromium is more toxic in soft water. Chronic chromium

toxicity to D. magna was reported at 66 μg/L in soft

water, whereas 44 μg/L inhibited reproduction in hard

water. Fish are more susceptible to inhibitory effects of

chromium at low pH. Table 5-24 presents data on toxicity

of chromium to freshwater organisms.

Sorensen (1991) states that maximum selenium levels

for fish should not exceed 3 to 8 μg/L. Inorganic selenate

has been shown to be acutely toxic to aquatic organisms

at levels as low as 760 μg/L (U.S. EPA, 1986c). HDR

(1995) indicated that no chronic toxicity data were available on selenium effects on freshwater organisms. Toxicity effects of selenium vary according to environmental

conditions such as differences in oxygen tension, habitat, food availability, trophic level of fish, chemical form,

fish age, dietary consumption, population composition

and competition, presence of other metals, and thermal

gradient (Sorensen, 1991). A summary of selenium toxicity data is presented in Table 5-25.

Table 5-22. Toxicity of Arsenic to Freshwater Organisms (HDR, 1995)

Species

Life

Stage

Hardness 

(mg/L) pH (SU)

Concentration

(mg/L)

Duration

(days) Effect

Oncorhyncus mykiss Eggs — — 0.1 180 No effect migration

Adult — — 13.3 6 LC50

2 months 13.2 8.3 1 21 Reduced growth

Parr 385 8 13.2–19.4 6 LC50

Adult 45 7.3 <0.09 28 No mortality

Ictalurus lacustris Adult — — 3,050 4 LC50

Lepomis macrochirus Immature 310 7.6 0.2–0.7 112 Growth unaffected

Adult 310 7.6 0.02–2.3 112 Growth/survival unaffected

Micropterus dolomieu Fingerling — — 900 4 LC50

Pimephales promelas Fry 47 7.2 5–19 1–4 No mortality

Egg 47 7.8 <16.5 29 No effect hatching

Egg 47 7.8 <2.13 29 No effect survival/growth

4 cm 48 7.2 135 4 LC50

Crustacea

Bosmina longirostris 1 day 120 6.8 0.85 4 EC50

Cyclops (mixed spp) Nauplii 139 7.6–8.8 1.6–10 14 <20% mortality

Daphnia magna — — 0.8–31 1–4 LC50

1 day 47 7.2–8.1 1.32 28 50% mortality, no young,

growth reduced 10%

1 day 47 7.2–8.1 0.6–1.8 1–4 LC50

Daphnia pulex 1 day 120 6.1–7.8 49.6 2 EC50

Gammarus

pseudolimnaeus

Adult 47 7.2–8.1 0.8–2.0 2–4 LC50

Adult 44 6.9–7.3 0.09–0.96 14 10–100% mortality

Mollusca

Aplexa hyorumpn Adult 49.5 7.4–7.7 24.5 4 LC50

Helisoma companulata Adult 44 6.9–7.3 0.85–0.97 28 No mortality

Insecta

Pieronarcys dorsata Larva 44 6.9–7.3 0.85–0.97 28 No mortality

Tenytarsus dissimilis Larva 47 7.2–7.7 97 2 LC50

Rotifera

Philodina roseola — — 5.0–18 4 LC50

Benthic fauna (mixed) — 0.5–3.0 112 50–90% reduction in

numbers

Key

LC50 = Lethal concentration of substance causing mortality in 50 percent of the population exposed to the substance.

95
Table 5-23. Toxicity of Cadmium to Freshwater Organisms (HDR, 1995)

Species

Life

Stage

Hardness

(mg/L) pH (SU)

Concentration

(mg/L)

Duration

(days) Effect

Oncorhyncus mykiss Eggs 80 7.9 >0.124 13 No effect migration

Alevins 80 7.9 >0.124 27 LC50

Fry 80 7.2 0.091–0.677 2 Reduced growth

Juvenile 125 7.8 0.01–0.03 7–10 LC50

Adult 82 7.8 0.006–0.40 4–14 No mortality

Castostomus commersoni Adult 18 6.3 0.95–6.7 1–4 LC50

Embryo — — >0.012 Reduced growth

Cyprinus carpio Adult 55 8 0.24–0.45 1–4 LC50

Lepomis macrochirus Yearling 200 7.2 >0.08 330 90% survival

Yearling 200 7.2 0.757–2.14 270 100% mortality 

Egg/Larvae 200 7.2 0.031–0.80 330 No effect

Adult 200 7.2 >0.757 6 No spawning

Larvae 200 7.2 0.031 60 18% mortality

Larvae 200 7.2 0.08 30 70% mortality

Pimephales promelas Immature 201 7.2 <0.057 30 No effect survival/growth

Immature 201 7.2 0.068–0.150 270 LC50

Fry 204 7.6 <0.057 30 No effect survival/growth

Eggs 204 7.6 <0.027 180 Hatchability

Eggs 204 7.6 <0.037 180 Spawning unaffected

Crustacea

Cyclops abyssorum Adult 50 7.2 3.8 2 LC50

Daphnia magna 1 day 130 6.95 0.005–0.008 2–4 LC50

Mature — 7.8–8.6 >0.5 1–3 100% mortality, <25°C

Female — 7.8–8.6 <0.025 1–3 No mortality, @ 30°C

Daphnia pulex 1 day 106 8.5–8.6 0.115 1 LC50

0.015 70 Brood size reduced

0.005 58 Survival reduced

Eudiaptomus padanus Adult 50 7.2 0.55 2 LC50

Gammarus pulex Adult 40 — 0.12–0.68 2–4 LC50

Cyclops (mixed sp.) Nauplii 139 7.6–8.8 0.04 14 100% survival

0.016 14 54% mortality

0.4 14 73% mortality

Annelida

Tubifex tubifex Adult 224 8.5 450 1 LC50, 28°C

224 7 320 2 LC50, 10°C

Limnodrilus hoffmeisteri Adult 224 7 0.2–0.55 4 LC50

Mollusca

Physa integra Adult 46 7.1–7.7 0.02–0.114 7–21 LC50

Lymnaea stagnalis Adult soft — 0.11–0.33 7–14 LC50 

Insecta

Ephemerella subvaria Larva 54 7 2 4 LC50

Ephemerella spp. Larva 46 7.1–7.7 0.085 21 No mortality

Hydropsyche betteni Larva 46 7.1–7.7 0.238 28 No mortality

Chironomus tendipes Larva 220 7 25 2 LC50

Key

LC50 = Lethal concentration of substance causing mortality in 50 percent of the population exposed to the substance.

96
Mercury (Hg) can accumulate in tissue; concentrations

as high as 27.8 mg/L have been measured in fish muscle (Sorensen, 1991). In general, mercury levels greater

than 1 μg/L have been toxic to freshwater organisms

(Table 5-26). Chronic mercury toxicity to fathead minnows in their early life stage has occurred at 0.26 μg/L.

The estimated mercury levels in the canal water column

were less than 1 μg/L and were generally less than 0.2

μg/L, except for Deer Valley during minimum flow conditions (0.5 μg Hg/L) (see Table 5-21).

Table 5-24. Toxicity of Chromium to Freshwater Organisms (HDR, 1995)

Species

Life

Stage

Hardness

(mg/L) pH (SU)

Concentration 

(mg/L)

Duration

(days) Effect

Oncorhyncus mykiss Eggs 80 7.8 0.02–2.0 9 Unaffected

Juvenile 80 7.8 2 224 50–100% mortality

Alevins 80 7.8 0.02–2.0 224 Unaffected growth/survival

Alevins 80 7.8 0.2–2.0 224 50–100% mortality

Fry 70 7.8 0.013 60 Reduced growth

Fry 70 — 0.017 60 94% mortality

Adult 70 — 0.077–0.170 60 Unaffected growth/survival

Adult — — 8.88–32 4 LC50

4–9

months

70 — 7.0–180 4 LC50

Fingerling 80 7.8 2.0–50 4 30–70% mortality

Castostomus commersoni Egg/Fry 35 — >0.29 60 Reduced growth

Cyprinus carpio Adult 55 8 14.3–21.2 1–4 LC50

Lepomis macrochirus Egg/Fry 35 — 1.122 60 Reduced growth

Adult 35 — <0.522 60 No effect

Pimephales promelas Immature 209 7.5 27–58 25 No effect survival/growth

11 weeks 209 7.5 18.0–140 1–11 LC50

Juvenile 209 7.5 1–3 400 No effect spawning/hatching

Adult 270 7.6 38–61 2–7 LC50

Crustacea

Cyclops abyssorum Adult 50 7.2 10 2 LC50

Daphnia magna 1 day 50 7.8 1.8 1 LC50

Eudiaptomus padanus Adult 50 7.2 10.1 2 LC50

Adult 50 7.2 0.55 2 LC50

Gammarus pseudolimnaeus Adult 45 — 0.067 4 LC50

Gammarus sp. Adult 50 — 3.2 4 LC50

Annelida

Nais sp. — 50 — 9.3 4 LC50

Mollusca

Lymnaea emarginata Adult 154 — 34.8 2 LC50

Physa integra Adult 154 — 0.66 2 LC50

Insecta

Ephemerella subvaria Larva 50 6 2 4 LC50

Hydropsyche betteni Larva 42 — 64 4 LC50

Tanytarsus dissimilis — 44 — 59.9 4 LC50

Rotifera

Philodina acuticornis — 25 — 3 4 LC50

— 81 — 15 4 LC50

Key

LC50 = Lethal concentration of substance causing mortality in 50 percent of the population exposed to the substance.

97
Combinations of dissolved metals in the aquatic environment may have antagonistic, synergistic, and additive effects on organisms (Esernick et al., 1991; Goss

and Wood, 1988; Spehar and Fiandt, 1986; Mukhopadhayay and Konar, 1985; Kaverkamp et al., 1983; Kaviraj

and Konar, 1983; Suffern, 1981; Westerman and Birge,

1978). These effects will vary with respect to environmental conditions and exposed indigenous organisms.

The State of Arizona and EPA acute and chronic water

quality standards for heavy metals to protect aquatic life

are presented in Table 5-27. Based on these standards

and the source water quality data presented in Table 5-10,

cadmium (6 to 7 μg/L) and mercury (0.04 to 0.26 μg/L)

levels in the Verde and Salt Rivers exceed chronic

standards. Contribution of cadmium (12 μg/L) from the

Deer Valley WTP may cause the levels in the canal to

exceed state and federal acute toxicity standards, 5.6

μg/L and 3.9 μg/L, respectively.

Suspended Solids

HDR (1995) states that sedimentation of residuals solids

may affect benthic communities, algae, aquatic macrophytes, and suitability of fish spawning habitat. The

European Fisheries Advisory Commission states that

salmonids will not spawn in areas of high solids deposition (EIFAC, 1969). Siltation shortly after spawning may

affect the proper transfer of oxygen and carbon dioxide

between the egg and adjacent water and, thus, result in

mortality (EIFAC, 1969; U.S. EPA, 1986c).

Deposition of solids can also smother benthic organisms

(Gammon, 1970; Tebo, 1965; George et al., 1991). Furthermore, alum sludge sediments may adversely affect

benthic macroinvertebrates by limiting the availability of

carbon (George et al., 1991). Deposition of solids can

also reduce species diversity of the benthic organisms

(Mackenthun, 1973; George et al., 1991). As previously

mentioned, the deposition of residuals solids depends

on stream velocity. George et al. (1991) found that

deposition rates could be reduced and, therefore, the

detrimental effects of alum sludge deposits to benthic

communities could be minimized by discharging residuals solids during periods of high stream velocities.

HDR (1995) concludes that any effects of WTP residuals

deposition on fish spawning would be considered negligible to the City, since the white amur, the most economically valuable species, is sterile and will not spawn

anyway. Arizona’s surface waters must comply with the

state Water Quality Standards for Navigable Waters,

which states that “navigable waters shall be free from

pollutants in amounts or combinations that settle to form

deposits that adversely affect aquatic life, impair recreational uses, or are unsightly.” Discharge of residuals

Table 5-25. Toxicity of Selenium to Freshwater Organisms (HDR, 1995)

Species

Life

Stage

Hardness

(mg/L) pH (SU)

Concentration 

(mg/L)

Duration

(days) Effect

Oncorhyncus mykiss Juvenile 135 7.9 5.41 4 LC50 daily feed

Uneyed eggs 0.47 and less 42 No mortality

Fingerlings 0.47 and less 42 No effect

Mortality/growth

Adult 36 12.5 4 LC50

Carassius auratus Juvenile 148 7.6 36.6 4 LC50

8.8 14 LC50

Cyprinus carpio 3–26 g — 7.2–8.2 35 4 LC50

Ictalurus punctatus Juvenile 140 7.9 19.1 4 LC50

Lepomis macrochirus Juvenile 150 7.75 30.7 7 LC50

17.6 14 LC50

Micropterus salmoides Juvenile 25 & 200 6.0 or 8.7 0.01 120 No mortality

Pimephales promelas Juvenile 151 7.8 7.3 4 LC50

2.9 7 LC50

Crustacea

Daphnia magna 1 day 140 8.2 1.87 7 LC50

1.5 32 Reproduction unaffected

1 32 Growth unaffected

Daphnia pulex 1 day 46 7.4 0.6 28 No effect

Adult 46 7.4 1.374 2 LC50

Key

LC50 = Lethal concentration of substance causing mortality in 50 percent of the population exposed to the substance.

98
solids may violate state surface water quality standards

by reducing benthic community populations and diversity, and destroying suitable spawning beds (HDR, 1995).

5.3.4.4 Conclusions

After evaluating the available data, HDR (1995) concludes that no adverse effects to aquatic organisms are

anticipated from metals (aluminum, arsenic, cadmium,

chromium, selenium, and mercury) contained in WTP

waste stream discharges. If low flow conditions (25

percent of typical flow) are maintained, cadmium concentrations in the canal downstream from the Deer Valley WTP may exceed state and EPA water quality

standards. The hardness, circumneutral pH, and high

alkalinity of the waters in the SRP canal system would

help reduce any potential inhibitory effects from metals

on aquatic life. Solids deposition may affect benthic

communities and eliminate spawning areas and the successful development of fish eggs. HDR (1995) further

states that the impacts may not be important unless the

Arizona Game and Fish Department attempt to create a

sustainable urban fishery in the canals.

Table 5-27. Maximum Contaminant Levels (μg/L) Cold Water

Fishery (HDR, 1995)

Acute Standard Chronic Standard

Contaminant State EPA State EPA

Arsenic (total) 360 360 190 190

Cadmium (total) 5.6a 3.9 1.5 1.1

Chromium III (total) 2,261 1,700 269 210

Mercury (dissolved) 2.4 2.4 0.01 0.012

Selenium (dissolved) 20.0 260 2.0 35

TTHM NNS 11,000 NNS NNS

a Value for warm water fishery is 34.6.

Key

NNS = No numerical standard.

Table 5-26. Toxicity of Mercury to Freshwater Organisms (HDR, 1995)

Species

Life

Stage

Hardness

(mg/L) pH (SU)

Concentration

(mg/L)

Duration

(days) Effect

Oncorhyncus mykiss 2 month 82–132 6.4–8.3 0.33 4 LC50

Carassius auratus Adult 100 — 0.35 2 LC50

Channa punctatus 55–g 160 7.4 1.8 4 LC50

Cyprinus carpio Adult 55 8 0.18 4 LC50

Lepomis gibbosus Adult 55 8 0.3 4 LC50

Lepomis macrochirus 0.6 g 46 7.1–7.3 0.16 4 LC50

Pimephales promelas 3.2–4.2 cm 40–48 7.2–7.9 1–Jan 4 LC50

Tilapia mossambica Adult 115 8.5 1 14 LC50

Annelida

Limnodrilus hoffmeisteri Adult 5 7 0.15–0.50 4 LC50

Tubifex tubifex Adult 5 7 0.14–0.27 4 LC50

Planaria

Dugesia dorotocephala Adult — — 0.2 10 no mortality

Crustacea

Cyclops (mixed sp.) Nauplii 139 7.6–8.8 <0.032 14 100% survival

Cyclops abyssorum Adult 50 7.2 2.2 2 LC50

Eudiaptomus padanus Adult 50 7.2 0.85 2 LC50

Mollusca

Aplexa hypnorum Adult 50 7.2–7.4 0.37 4 LC50

Insecta

Chironomus tendipes Larva 220 7 64 1 LC50

Ephemerella subvaria Larva 42 7.6 2 4 LC50

Hydropsyche betteni Larva 42 7.6 2 4 LC50

Key

LC50 = Lethal concentration of substance causing mortality in 50 percent of the population exposed to the substance.

99
5.4 Recommended Practices

Limited information exists regarding the impact of direct

discharge of WTP residuals solids on aquatic environments. Regulatory agencies have permitted the direct

discharge of residuals to streams and impoundments

because of the absence of information showing any adverse effects. Some states have taken a more conservative approach recently and have discontinued allowing

discharge of residuals solids to freshwater. The first step,

therefore, by a utility considering direct discharge of WTP

residuals into a receiving stream is to determine its state

regulatory requirements (see Figure 5-12).

The scientific literature and examples discussed in this

chapter show that deposition of residuals solids can

adversely affect benthic community populations and diversity. These direct consequences on benthic organisms may limit a food source for certain fish species. In

addition, solids deposition may affect fish egg survival.

Discharge of WTP residuals in sluggish streams or lakes

should be avoided. A utility can use hydrologic models

such as the U.S. Army Corps of Engineers’ HEC-2 or

HEC-6 models to estimate possible deposition within a

receiving stream. More frequent, reduced solids mass

loading from the WTP during high stream flow will possibly reduce deposition and adverse effects to downstream benthic organisms.

The receiving water chemistry can minimize toxic effects

of metals in the discharged waste stream. Hardness,

alkalinity, pH, dissolved oxygen, sulfate, and other water

quality parameters minimize the inhibitory effects of

heavy metals. Because of the potential toxicity to

aquatic organisms, utilities that are evaluating discharge

to soft waters (hardness less than 50 mg CaCO3/L)

should consider an alternate residuals disposal method.

Furthermore, receiving waters with a pH less than 6

Figure 5-12. Recommended practices for direct discharge of WTP residuals to surface waters.

100
should be avoided. Discharge to these waters could

result in increased solubility of metals and increased

toxic effects. The chemical characteristics of the waste

streams must be determined as well as the contaminant

mass loadings to the receiving stream. Using information on stream hydrology and stream and waste quality,

a mass balance should be conducted to estimate potential concentrations of contaminants downstream from a

prospective point of discharge. Consideration of mixing

zone should be included in modeling the system.

When considering the future use of a receiving water, a

utility should consider whether direct discharge of its

residuals will negatively affect the downstream aquatic

life or quality. As part of this consideration, the utility

should evaluate upstream and downstream water quality, sediments, and aquatic life to determine if the water

has already been negatively affected by some previous

use. To fully evaluate the toxicity of the WTPs discharged residuals, a series of toxicity tests should be

conducted on a variety of aquatic organisms that are

indicative of the organisms inhabiting the aquatic ecosystem. As shown in the case studies, one toxicity test

may not adequately describe toxic effects.

Residuals from WTPs using aluminum salts have

been the focus of research on the impacts of direct

discharge of wastestreams to freshwater aquatic ecosystems. The recommended protocol presented in

Figure 5-12, however, can be followed for any type of

discharge. Aluminum may not be inhibiting organism

growth or survival; other metals or contaminants contained in wasted residuals may be the primary toxicant. Detrimental effects of solids deposition on

benthic organisms can be caused by limiting oxygen

transfer to chemical toxicity. In any case, sedimentation of solids must be evaluated and anticipated.

Other chemical residuals can affect benthic communities and fish spawning areas. In addition, chemical

characteristics and quantities of the receiving water

and the wastestreams must be determined to anticipate chemical changes to the water downstream

from the points of discharge. Toxicity testing must

be conducted to determine potential toxic effects on

the ecosystem.

101
Chapter 6

Discharge to Wastewater Treatment Plants

Discharge of water treatment plant (WTP) residuals to a

wastewater treatment plant (WWTP) is an attractive

disposal option for water utilities. This chapter provides

information on the experience of utilities using this practice, factors that should be considered, regulatory limitations, and design considerations for conveyance

facilities (both gravity and pumped systems) and for the

receiving WWTPs. The impacts on options for ultimate

disposal of residuals from WWTPs are also considered.

6.1 Background

Many water utilities across the country discharge WTP

residuals to a sewer and/or to a WWTP. These options

are often economically attractive, and transfer disposal

liability to the WWTP. The American Water Works Association (AWWA) Water Industry Database (AWWA/

AWWARF, 1992), a survey of WTPs currently discharging to a WWTP, reveals that such plants are variable in

size, and typically treat surface water supplies as opposed to ground waters. Table 6-1 summarizes the facilities surveyed.

Conventional WTPs (coagulation, sedimentation, filtration) commonly discharge filter backwash solids and/or

clarification basin residuals to a sanitary sewer system

for eventual treatment at a WWTP.

Several factors must be considered when evaluating the

feasibility of discharging residuals to a WWTP. The interests and concerns of WWTP managers and operators

are different from those of similar personnel in water

utilities. Factors for a WWTP to consider are available

capacity (conveyance system and treatment plant),

treatment process compatibility, and final disposal requirements. Introducing water treatment residuals to a

WWTP may offer some benefits in terms of process

performance. On the water treatment facility side, pretreatment requirements, storage facilities, and conveyance systems must be considered. Costs and service

agreement terms must also be evaluated.

Regulatory requirements, especially those related to ultimate disposal, must also be addressed. Essentially, all

states allow for discharge of WTP residuals (sludge and

brines, or reject waters) to a WWTP, but the specific

requirements and limitations vary. Regulatory requirements imposed on WWTPs will be directly or indirectly

imposed on the WTP discharging to the WWTP. Quality

issues, especially levels of metals and other inorganics,

are likely to be a concern.

6.2 Survey of Operating Systems

The results of a nationwide survey of WTPs currently

discharging to WWTPs are shown in Appendix B.

6.3 Design Considerations and

Conveyance Systems

For WTP managers and operators, disposal of WTP

residuals into a sewer system or to a wastewater treatment facility is usually very attractive, cost effective, and

offers significant benefits from a regulatory and operational standpoint (Weaver, 1985; Price et al., 1989;

Robertson and Lin, 1981; Reh, 1978). This disposal

method, however, has not been universally accepted by

the wastewater community (Novak, 1989). Provided that

the water treatment facility complies with the receiving

wastewater utility’s pretreatment and discharge requirements, the liability for proper disposal of the water treatment residuals is transferred from the water utility to the

wastewater utility. Operationally, a WTP has no further

responsibility for its waste residuals once they have

been transferred to a wastewater utility; routine maintenance of the sewer system and/or operation of the

wastewater treatment facility is the responsibility of the

wastewater utility.

Costs associated with this option include the capital

expenditures necessary to intercept, tie in, or transport

residuals; user fees imposed by the receiving wastewater utility to recover conveyance; treatment and ultimate

disposal operation and maintenance (O&M) costs; and

other costs. These costs are often less than other options for directly handling WTP residuals. In some

cases, the receiving WWTP benefits from increased

removal of suspended solids and/or biochemical/chemical oxygen demand (BOD/COD), additional phosphorus, and H2S conversion at the treatment plant (van

Nieuwenhuyze et al., 1990).

WWTP operators, however, may be increasingly

reluctant to accept WTP residuals because of

102
concerns regarding National Pollutant Discharge

Elimination System (NPDES) compliance and toxicity.

For example, the Charlotte-Mecklenburg Utility Department will limit future sewer disposal of residuals from

its Franklin WTP because this practice violates pretreatment ordinances (Appendix B). Potential limiting

concerns for a WWTP when it accepts WTP residuals

include increased final suspended solids, decreased

effective digester capacity, overloading of primary

clarifier and sludge removal systems, and overloading

of dewatering operations (van Nieuwenhuyze et al.,

1990).

Table 6-1. Survey of Water Treatment Plants Discharging to WWTPs (AWWA/AWWARF, 1992)

Size

Facility

Avg. Daily

(mgd)

Design

(mgd) Major Processes

Dunkirk Water Treatment Plant

Dunkirk, NY

4 8 Coagulation, lime softening, filtration

City of Boulder WTP

Boulder, CO

5.3 12 Coagulation, sedimentation, filtration

Chesterfield County WTP 10.5 12 Fe/Mn removal, coagulation, lime softening, filtration

Conestoga WTP

Lancaster, PA

7 16 Coagulation, sedimentation, filtration

T.W. Moses WTP

Indianapolis, IN

10 16 Coagulation, sedimentation, filtration

Texarkana WTP

Texarkana, TX

11 18 Sedimentation, filtration

Franklin WTP, CharlotteMecklenburg Utility District47.6 96 Coagulation, sedimentation, filtration

Vest WTP, Charlotte-Mecklenburg Utility

District

18.9 24 Coagulation, sedimentation, filtration

City of Myrtle Beach WTP

Myrtle Beach, NC

14 29.5 Coagulation, sedimentation, filtration, ozone

City of Greensboro WTP

Greensboro, NC

31 44 Coagulation, sedimentation, filtration

Knoxville WTP

Knoxville, TN

30-33 60 Coagulation, sedimentation, filtration

Erie City Water Authority

Erie, PA

35-40 60-80 Coagulation, sedimentation, filtration

Chattanooga WTP

Chattanooga, TN

38 72 Coagulation, sedimentation, filtration, GAC

Belmont WTP

Philadelphia, PA

60 80 Coagulation, sedimentation, filtration

Queens Lane WTP

Philadelphia, PA

100 100 Coagulation, sedimentation, filtration

Nottingham WTP

Cleveland, OH

87.1 100 Coagulation, sedimentation, filtration

Morgan WTP

Cleveland, OH

80 150 Coagulation, sedimentation, filtration

Baldwin WTP

Cleveland, OH

85 165 Coagulation, sedimentation, filtration

Nashville WTP

Nashville, TN

120 180 Coagulation, softening, filtration

City of Milwaukee Waterworks

Milwaukee, WI

60 Winter, 120

Summer

250 Coagulation, sedimentation, filtration

F.E. Weymouth Filtration Plant

LaVerne, CA

330 520 Coagulation, sedimentation, filtration

Robert A. Skinner Filtration Plant

Temecula, CA

300 520 Coagulation, sedimentation, filtration

Joseph Jensen Filtration Plant

Granada Hills, CA

333 550 Coagulation, sedimentation, filtration

103
Discharge to a sewer is possible only when water treatment and wastewater treatment facilities work together

cooperatively. Information sharing is imperative to establish such a relationship. A water utility should provide

information concerning residuals quality and quantity,

dry solids content, pH, nutrients, heavy metals, and

other relevant parameters. In addition, a wastewater

utility and water utility must cooperate to determine

whether positive or negative effects can be expected at

the wastewater facility with respect to hydraulic conditions in the sewer lines, hydraulic and process capacity

at the WWTP, sludge treatment, and ultimate disposal

of WWTP biosolids.

6.3.1 Regulatory Considerations

Regulations governing discharge to WWTPs are discussed in Chapter 2. Hazardous waste regulations (federal and state) are usually not a problem, since WTP

residuals are rarely, if ever, classified as hazardous

waste. Radioactive waste regulations are also not a

problem, since the radioactive component of WTP residuals usually is very low and results from the removal

of normally occurring radioactive material (NORM) from

the raw water supply. Local receiving wastewater regulations are usually driven by the need for a utility to

comply with provisions of the Clean Water Act. These

local regulations are imposed to reduce the risk of operational problems at the receiving plant or of violation

of its discharge permits.

Discharges to sanitary sewers are subject to the U.S.

Environmental Protection Agency’s (EPA’s) National

Pretreatment Standards and possibly to more stringent

pretreatment requirements imposed by the state or

WWTP. The requirements imposed by a wastewater

treatment facility are necessary to enable the facility to

achieve compliance with its NPDES permit (including

provision for sludge disposal). Pretreatment standards

are typically industry or site specific. The local discharge

requirements imposed by a wastewater utility may be

governed by:

• Impact of residuals on the waste conveyance system

(sewer, pump station, force mains, abrasion, and corrosion).

• Impact of residuals on either the liquid or solids process treatment system at the wastewater treatment

facility (e.g., rate of discharge, solids concentrations,

need for flow equalization).

• Concerns about biotoxicity (within the WWTP or in

the plant’s effluent).

The wastewater facility may refuse to accept the waste,

or it may impose extremely strict local limits on the

supplier’s discharge (Koorse, 1993b).

Limitations often placed on WTP residuals discharged

to sanitary sewers and/or to WWTPs are the result of

problems perceived or experienced by the receiving

wastewater treatment facilities. The data in Appendix B

indicate that discharge of residuals through a force main

from a WTP in Knoxville, TN, caused a high degree of

solids accumulation in flocculation basins and considerable yearly maintenance. In a study of 13 WTPs operated by the American Water Works Service Company in

West Virginia (EET, 1992), five of the plants discharged

directly to sanitary sewers, and three plants hauled their

residuals to a WWTP. Sewer systems at two locations

experienced blockages that resulted in backwash water

and WTP residual overflows to surface water. Only one

plant is operating under a local industrial user permit,

stipulating maximum heavy metal concentrations, pH

and flow limitations, and monitoring requirements.

As greater concern is paid to WWTP effluent toxicity,

municipalities are adopting pretreatment regulations to

limit the strength or quality of discharge to sanitary

sewers. Aquatic toxicity caused by aluminum in wastewater plant effluent has been widely studied. Hall and

Hall (1989) report that a substantial increase in the

mortality rate of Ceriodaphnia dubia is not observed in

100 percent alum effluent, but delayed brood release

and significant reduction in reproduction are observed.

Reductions in pH and dissolved oxygen concentrations,

high levels of suspended solids, and, possibly, aqueous

aluminum were implicated as likely causes of toxicity. In

Norfolk, Virginia, WWTP effluent toxicity was tested

(Tsang and Hurdle, 1991). When the greatest quantities

of alum sludge were fed into the wastewater plant influent, no acute but some chronic toxicity results were

measured. At lower feed rates of alum residuals to the

plant, no toxic effects were noted.

Larger WTPs that discharge their residuals to a sanitary

sewer or WWTP are usually regulated by local industrial

user permits. Examples of this are Indianapolis, Milwaukee, and several large WTPs at the Metropolitan Water

District in southern California, where a surcharge is

required if heavy metal limits are exceeded (see Appendix B).

The City of Philadelphia’s Queens Lane and Belmont

WTPs discharge all of their residuals to sanitary sewers

and operate under combined effluent discharge permits

required in-house between the water and wastewater

plants. The permits are administered by the Industrial

Waste Unit of the Philadelphia Water Department. Limitations were placed on the discharge from these plants

because of NPDES permit violations by the WWTPs.

The City of Philadelphia’s permits limit the discharges

resulting from periodic raw water basin cleaning, routine

flocculation/sedimentation tank discharges, and periodic flocculation/sedimentation tank cleaning. The user

permit requires notification of discharges, sampling and

104
monitoring, onsite detention for equalization and solids

relocation, prohibition of accidental or slug discharges,

and maximum 3-day discharge levels (lb/day dry weight)

of residuals solids that may be discharged (AWWA/

AWWARF, 1992).1

Municipal discharge permit ordinances are written to

protect the receiving WWTPs and are industry oriented.

The limitations specified are often too stringent for the

discharge of WTP residuals. When the discharge of

WTP residuals is allowed, discharge permit limitations

may include:

• Limitations on maximum discharge rates, often during

stated periods of the day.

• Maximum BOD concentration limits.

• Maximum total suspended solids (TSS) concentration

limits.

• Total dry solids limits—tons/day maximum.

• Cessation of discharge upon notification during high

storm flow periods.

• Prenotification of discharge occurrences.

6.3.2 Conveyance System Design

Considerations

6.3.2.1 Selection of Type

Three types of conveyance systems generally are used

to transport WTP residuals to a wastewater treatment

facility: 1) gravity sanitary sewers, 2) pumping/force

main systems, and 3) truck transport. Factors that affect

the choice of system include the receiving wastewater

system’s availability and capacity, impacts of the residuals on the treatment plant, cost, corrosion considerations, and ultimate disposal considerations. Discharge

to sanitary sewers, where available, is often the least

expensive and preferred choice for a WTP. This option,

however, provides the receiving wastewater utility with

the least treatment flexibility—all of the WTP residuals

must be processed through the wastewater liquid

stream processes. If a WWTP has excess design capacity beyond its normal wastewater loading, WTP residuals received through the sanitary sewer generally

can be also processed.

When problems develop because of WTP residuals,

they usually occur in the liquid stream processes. Alternatives to the sanitary sewer are generally more costly:

discharge by pumping/force main or truck haul directly

to the wastewater treatment facility. With either of these

alternatives, the discharge is controllable and can be

directed to the desired liquids or solids stream process

at the WWTP. The WTP residuals do not have to go

through the liquid stream treatment process at all. They

can either be directed through the solids processing

facilities, in a combined facility or separately, or they can

be combined with the wastewater treatment biosolids

after the solids processing facilities for ultimate disposal.

There are other reasons for considering alternatives to

the sanitary sewer for conveying WTP residuals to a

wastewater plant. Koplish and Watson (1991), for example, described a situation in Allentown, Pennsylvania,

where one entity is responsible for all water and wastewater functions. There the preferred, most cost-effective

option for dewatering residuals (water and wastewater)

is to have all equipment at a single location. This, in turn,

necessitates direct transport (versus gravity sewer

transport) of the alum residuals to the wastewater treatment facilities, where they are fed into a thickener prior

to belt filter press dewatering.

6.3.2.2 Pretreatment Requirements

Wastewater utilities often impose requirements on a

water treatment facility that govern the release of residuals to a sanitary sewer. The most common of these

requirements is equalization of discharge flow. Other

pretreatment requirements can include regulating the

quality of the discharge, which may include pH neutralization; homogenization of the waste stream to ensure

uniform concentration versus slug concentration discharges; limits on the total solids allowed to be discharged; and limitations on quality parameters such as

heavy metals or components that may cause corrosion,

odors, or other undesirable conditions.

Equalization

Directly discharging WTP residuals to a sewer system

in a fairly continuous and uniform way may be possible

if the sedimentation basins at a WTP are equipped with

residuals removal mechanisms. Generally, flow equalization is required to ensure a uniform sludge flow to the

sewer. Equalization facilities provide storage for quantities of waste discharge that exceed the allowable discharge to a sewer system. The storage requirements

depend on the designated waste discharge schedule.

Flow equalization is almost always required when sedimentation basins are being cleaned, to control wastes

from ion exchange processes that are only produced

during media regeneration, and before discharging filter

backwash water to a sewer.

The need to control the rate of water treatment waste

discharge into a sewer system was shown in the 1989

Durham study by McTigue et al. (1989). The water treatment facility examined had sedimentation basins in two

groups of four basins each. Each basin had six conical

hoppers in the inlet end; each hopper was approximately

1 Wankoff, W. 1992. Personal communication between C.P. Houck

and W. Wankoff, Water Treatment Plant Manager, City of Philadelphia Water Department. December 14.

105
10 feet deep and had a drain in the bottom for the

removal of residuals. Normal operation of the plant consisted of flushing the residuals from each hopper daily.

Half of the basin hoppers were discharged to a sanitary

sewer for 2 weeks, with no significant effect on the

receiving WWTP effluent quality.

Based on these results, the remaining hoppers were

connected so that all hopper-flushing residuals could be

sent to the wastewater plant. Only slight increases in

effluent suspended solids, turbidity, and color were observed from the additional hopper flushings. Problems

did occur, however, when a complete basin was cleaned

and the residuals were discharged over 6 hours. A slug

discharge caused severe deterioration of the wastewater utility’s effluent quality in terms of turbidity, color, and

suspended solids. It took approximately 6 days for the

wastewater treatment facility to recover from the shock

load. In addition, the inert WTP solids captured by the

WWTP began to accumulate in the digester when the

operators failed to compensate by increasing the

amount of digested biosolids removed for ultimate disposal. The result was a deterioration in digester supernatant quality.

The need for equalization capabilities is often realized

during an unanticipated experience such as a situation

that developed in Philadelphia in 1988.2

 Philadelphia

has three WTPs, two of which—Queens Lane WTP (100

mgd) and Belmont WTP (60 mgd)—discharge their residuals to sanitary sewers. In 1988, the Queens Lane

WTP cleaned flocculation/sedimentation basins that are

stacked, with chain and flight residuals collectors on the

top, and no residuals collection on the bottom. A significant time had elapsed since the last basin cleaning. A

slug of approximately 400 tons (dry weight) of solids was

discharged into the sanitary sewer that then completely

overloaded the 210 mgd Southwest Water Pollution

Control Plant (then loaded at 120 mgd). Residuals from

the Southwest Plant were discharged via a 5-mile force

main to the Sludge Recycling Center, and so the additional 400 tons of WTP residuals were pumped as

quickly as possible into the Sludge Recycling Center,

overloading it as well.

The result of the poor coordination between the water

and wastewater treatment plants was that two WWTPs

fell out of compliance with their discharge permits. When

the permits were renewed, EPA required that an internal

city permit be issued by the Wastewater Department to

the Water Department, regulating the WTP residual discharges. Extensive monitoring was also required.

In Philadelphia, effluent limits have been established for

WTP discharges from cleaning of raw water basins,

routine flocculation/sedimentation tank discharges, and

periodic flocculation/sedimentation tank cleaning. Because their solids content is low, routine filter backwash

discharges are not regulated by permit, even though a

high volume of water is discharged. Flow, total suspended solids (TSS), aluminum, iron, and arsenic are

monitored on a daily maximum and monthly average

basis. TSS limitations of 132,000 lb/day (daily max) and

53,000 lb/day (monthly average) have been placed on

the Queens Lane plant discharges.

Similar limits are placed on Philadelphia’s Belmont plant

discharges. Total discharge to the city sewers must be

at a controlled rate such that no more than 10 percent

of the total solids in the basin is removed per day. TSS

concentrations must be below 100 ppm. Total allowed

maximum flow from either plant is 18 mgd, and it must

occur during the first shift only. Cleaning and discharge

from both plants cannot occur simultaneously. Additionally, the Queens Lane plant is required to maintain, in

good working order, an onsite lagoon to receive diversional flow streams. The need for the installation and

operation of equalization facilities for waste WTP residuals in Philadelphia is obvious, given its 1988 experience.

Equalization restrictions may include limitations on the

time of day of discharge, the maximum flow over a

certain period, or the maximum solids discharge that is

allowed. Cornwell et al. (1987) present a mass storage

diagram approach that analyzes storage volume required for a given solids concentrations. The resulting

equalization basin is usually sized to be capable of

decanting and thickening. Either a continuous flow or

batch fill and draw thickener may be designed. The

capability to mix the contents of the equalization/thickener tank also may be desirable. Cornwell et al. (1987)

also discuss design considerations for equalization facilities to handle backwash water or similar high-flow,

short-duration discharges before their release to sanitary sewers. Sometimes, discharges of peak flows during otherwise low flow periods in the sanitary sewer may

be desirable.

Quality Discharge Limitations

Limitations may be imposed on TSS or total dissolved

solids (TDS), pH range, heavy metal constituents (either

from a receiving water quality or residuals standpoint, or

from a biotoxicity standpoint), and on corrosion causing

constituents. These limitations may be in a permit that

regulates discharge directly to the WWTP through a

force main or truck discharge, or in a permit for discharge to a sanitary sewer. Most municipal permit limitations imposed on industrial discharge to sanitary

sewer systems contain a provision that the water discharged must be in a pH range of 6 to 9. This could

cause problems for lime softening WTPs that either

have continuous discharge of softening residuals, or

periodic cleaning and discharge of sedimentation ba- 2 See footnote 1.

106
sins. Neutralization facilities for pH adjustment before

discharge at the water treatment facility may be required.

Heavy metals in the water environment are increasingly

regulated to ensure against toxicity to humans, plants,

and aquatic organisms. Heavy metals are generally concentrated in the residual solids of water treatment and

wastewater treatment wastes, and limitations on the

ultimate disposal of those residuals, as well as on the

dissolved fraction of heavy metals in the effluent, are

becoming more stringent.

Under the final City of Norfolk, Virginia/Hampton Roads

Sanitation District (HRSD) implementation plan to treat

WTP residuals at a district wastewater plant, WTP residuals were to be pumped directly to the wastewater

solids processing facility to circumvent problems both

with sludge settling and effluent toxicity (Tsang and Hurdle, 1991). The heavy metal content of the water plant

residuals, compared with HRSD’s Industrial Waste Discharge Regulations, indicated that the concentrations of

copper, chromium, lead, and zinc were close to the

HRSD limits (CDM, 1989).

In Philadelphia, the occurrence and level of arsenic in

the city’s WTP residuals discharged to the sanitary

sewer may be of concern.3

 The city’s Sludge Recycling

Center biosolids (received from the city’s WWTPs) have

an arsenic concentration of 11 to 12 mg/kg. The city’s

WTP residuals contribute less than 30 percent of the

total arsenic in the biosolids and are responsible for no

more than 2 to 3 mg/kg of the total. EPA’s 40 CFR Part

503 biosolids regulations establish a maximum allowable concentration of 41 mg/kg of arsenic in bulk sewage sludge, applied to a lawn or home garden, or in

sludge that is sold or given away. For land application

of biosolids, the State of New Jersey has set a maximum

limit of 10 mg/kg (New Jersey, 1993). The concentration

of arsenic in Philadelphia’s biosolids is not a problem

under the EPA regulation, but under New Jersey state

regulations, it may be. The city permit requires monitoring of the arsenic concentration discharged by the

WTPs to the city’s sanitary sewer system.

A similar situation exists in Massachusetts, where the

Type 1 cadmium and nickel standards of the Massachusetts Department of Environmental Protection are substantially more stringent than the federal 503 regulations

for “exceptional quality” biosolids (Donovan and Toffey,

1993). Should the arsenic levels become problematic, a

higher grade aluminum sulfate could be used which

generally contains lower levels of contaminants such as

arsenic.

Additional quality concerns that a WTP and a WWTP

receiving industrial discharge must consider are the potential corrosive nature of the WTP residuals and the

potential for hydrogen sulfide (H2S) generation. Aside

from the potential odor problem, the presence of H2S

can lead to corrosion of certain construction materials in

the sewer system. If the waste residuals to be discharged are septic (e.g., a lagoon cleaning operation) or

contain a high sulfate content, the sanitary sewer grades

are especially flat, or the sewers are constructed with

concrete pipe, the potential for H2S generation and corrosion attack should be investigated.

The discharge of WTP ferric residuals can counteract

the problems of H2S formation in sewers. Iron residuals

promote H2S binding. Iron residuals can also benefit the

wastewater treatment process through partial promotion

of presedimentation and partial phosphorus removal.

They can also minimize problems with sulfide or H2S

formation by keeping the H2S content in anaerobic digester gas at an acceptable level (van Nieuwenhuyze et

al., 1990).

The corrosive nature of brines resulting from ion exchange (IX), reverse osmosis (RO), or activated alumina

treatment of ground water is well known. Brines discharged to WWTP could be corrosive to piping systems

(Snoeyink et al., 1989) and may also upset the biological

balance in biological wastewater treatment systems.

A final quality concern is the amount of heavy metal

impurities in the water treatment chemicals used at the

WTP. The effect of contaminants on the chemical character of treated residuals can be significant (Lee et al.,

1990). The data in Table 6-2 analyze sludge from a

water treatment facility in Pennsylvania. Theoretical calculations show that the treatment chemicals could be

responsible for 100 percent of the levels of chromium,

copper, and lead detected in the residuals. A large fraction (78 percent) of the zinc concentration could also be

attributed to chemical addition. At the plant represented

in Table 6-2, the levels of zinc in the residuals prohibit

use for agricultural land application.

6.3.2.3 Gravity Sewers

Although WTP residuals will settle in the sanitary sewer,

their deposition should not occur more rapidly than sanitary solids (except residuals from lime softening and

diatomaceous earth WTPs). Sanitary sewers designed

with adequate slopes for conveying sanitary solids

should also convey WTP residuals. Generally, a velocity

of approximately 2.5 ft/sec (0.8 m/sec) or a residuals dry

solids content of less than 3 percent should be maintained to prevent sedimentation of hydroxide residuals

solids (van Nieuwenhuyze et al., 1990; Cornwell et al.,

1987). While not reporting the percent solids of its residuals, the Indianapolis Water Company did report that

its residuals were too thick for gravity flow (Appendix B).

Lime residuals may have settling velocities much higher

than those of metal coagulant residuals, and deposition 3 See footnote 1.

107
in sewer lines can be difficult to prevent. Robertson and

Lin (1981) recommend against the disposal of softening

basin residuals by direct discharge to a wastewater

utility because of the adverse effects on the sewer system conveying the wastes and on the equipment and

unit processes of the pollution control plant.

The capacity of the sanitary sewer system into which

WTP residuals are to be discharged should be checked

against the planned rate of residuals discharge and the

background flow in the sanitary sewer. Tsang and Hurdle

(1991) identify other considerations regarding conveyance that should also be checked:

• Ensure that there are adequate upstream dischargers

to flush the water plant residuals through the sewer

system.

• If there are no dischargers or too few dischargers

upstream, consider a discharge of “clean” filter backwash water after the residuals are discharged.

• If dilution flows are inadequate, plan on increased

maintenance at intermediate raw sewage pump stations to guard against damage from abrasive material

in the water plant residuals (especially if the raw

water source is a river) and from the low pH of the

residuals.

6.3.2.4 Pumping/Force Mains

Considerable data are published on the pumping of

water treatment residuals (e.g., Gandhi, 1992), as well

as on wastewater sludges (U.S. EPA, 1979b; Metcalf &

Eddy, 1991; Mulbarger et al., 1981). Relatively little

information is available, however, about the differences

in pumping dilute water treatment coagulant residuals

(less than 3 percent solids, dry weight) versus pumping

water. Dilute concentrations of coagulant residuals may

be pumped using centrifugal pumps with nonclog impellers (ASCE/AWWA, 1990). An operating point should be

selected to the right of the pump curve so that when the

pipe begins to clog, adequate head is available to unclog

it. A pump with a steep pump curve should be selected

to give a large reserve head. Precautions are required

to protect pumps and pipeline materials from corrosion

and abrasion, especially the abrasion of pump impellers.

If impeller abrasion is a serious concern, it may be

possible to use pneumatic ejectors to transport WTP

residuals (Foster, 1975).

In the design of a force main, the pipeline should be laid

out like a sewer—with constant slope—so pockets of

residuals do not solidify in the dips during periods when

pumping does not occur. Air release valves should be

provided at high points, but without using small orifices

that can clog. The pipeline should be provided with

flushing ports and cleanouts, and a drain facility at either

end so that the line can be flushed without creating a

problem. Pigging has been used for long reaches where

cleanouts are impractical, but rodding and jetting facilities are the most cost-effective cleaning alternatives for

shorter runs. The pipeline should be designed with a

minimum velocity of 2 ft/sec. If valving in the pipeline is

required, eccentric plug valves are best, but ball valves,

butterfly valves, pinch valves, and knife gate valves also

Table 6-2. Treatment Chemical Analysis Range of Detected Contaminant Levels (Dixon et al., 1988)

Contaminant

Ferrica

Chloride

(mg/kg)

Liquidb

Alum

(mg/kg)

Granular 

Alum

(mg/kg)

Percol

(mg/kg)

Cat

Floc-T

(mg/kg)

Cat

FlocTL

(mg/kg)

Hydrated

Lime

(mg/kg)

Pebble

Lime

(mg/kg)

Carbon 

(mg/kg)

Hydrofluosilicic

Acid

(mg/L)

Caustic

Soda

(mg/L)

C-9

(mg/L)

C-39

(mg/kg)

Arsenic (As) 108–122 214–270 555–621 <5.0 13 14 <5.0–15.5 <5.0 <5.0 3.4–18 <0.50 27 25

Barium (Ba) <5.0–7.4 <5.0 <50 <50 1.3 <5.0 24–83 <50 45–78 0.8–8.7 <5.0 1.7 <50

Cadmium (Cd) <0.5–59 <0.5–7.5 <5.0 <5.0 <0.5 <0.5 <5.0 <5.0 <5.0 1.8–2.9 <0.50 4.6 <5.0

Chromium (Cr) 148–568 16–24 <5.0 <5.0 <0.5 <0.5 <5.0–24 <5.0 <5.0 <0.50–2.9 <0.50 27.5 26

Copper (Cu) 32–840 <0.2–5.3 <2.0 <2.0 <0.2 1.3 <2.0–6.1 <2.0 5.9–13.5 <0.2–0.5 <0.20 1.4 <2.0

Lead (Pb) <0.5–332 <0.5 <5.0 <5.0 <0.5 2.8 <5.0–21 <5.0 <5.0 44–55 <0.50 52 44

Nickel (Ni) 1.5–114 <5.0 <50 <50 <5.0 2.2 <50 <50 <50 <5.0 <5.0 21 <50

Selenium (Se) 129–150 84–104 <5.0–204 <5.0 <0.5 10.7 <5.0–50 <5.0 <5.0 14.7–13.5 <0.50 39 35

Zinc (Zn) 69–7,990 <0.5–2.3 <5.0–12 <5.0 <0.5 1.6 <5.0–9.5 <5.0 <5.0–8.0 <0.5–1,418 <0.50 203,400 114,600

No. of Samples 8 6 3 1 1 1 4 1 3 4 21 1

a Solution 30 percent w/w; specific gravity = 1.346 g/mL. Contaminant concentration expressed in milligrams per kilogram (mg/kg) ferric chloride

solution. b Solution 8 percent w/w Al2O3; specific gravity = 1.33 g/mL. Contaminant concentration expressed in mg/kg liquid alum.

108
work. Using a valve that will not obstruct solids or rodding equipment is best.

EPA’s Process Design Manual for Sludge Treatment and

Disposal (U.S. EPA, 1979b) discusses the differences

between pumping water and wastewater residuals. In

water piping, flow is usually turbulent. Formulas for friction loss with clean water (Hazen-Williams, Darcy-Weisbach) are based on turbulent flow. Residuals flow may

also be turbulent, in which case the friction loss may be

roughly that of water. Residuals flow, however, is unlike

clean water in that laminar flow is also common. When

laminar flow occurs, the friction loss may be much

greater than that for water. Residuals up to 3 percent

solids concentration may have friction loss characteristics up to twice that of water under laminar flow

conditions when velocities are between 2.5 and 8 ft/sec.

The thixotropic behavior of residuals flow may significantly increase friction losses in residuals pumping, especially when restarting a pipeline that has been shut

down over a period of time. Head loss calculations must

consider the residuals viscosity and density, or alternatively, allow some safety factor of two to three times the

head loss with clean water. One rule of thumb for residuals thicker than 5 percent is to multiply the water head

loss by the percent total solids. In Pipeline Friction

Losses for Wastewater Sludges, Mulbarger et al. (1981)

prepared an extensive review of the literature on wastewater residuals pumping and rheology, resulting in a

methodology for predicting head loss relationships with

appropriate use of safety factors. Their document provides a series of predictive head loss curves for cast or

ductile iron pipelines, 4 through 20 inches in diameter.

Curves are presented for routine operation and worst

case (50 percent greater head loss than associated with

water) scenarios.

If the range of solids to be pumped is over 3 percent (dry

weight), the solids tend not to separate (especially over

6 percent) unless a super polymer or chemical (H2O2 or

acid) is added. In this event, pumping velocity is not so

important, since keeping solids in suspension is not a

serious concern. Also, head loss may preclude high

velocities. Positive displacement pumps are recommended (and are mandatory for solids greater than 4

percent). Cleanouts and flushing ports are mandatory.

Keeping the suction line short into the pump is also

important.

Communities in which WTP residuals are currently

pumped include Norfolk, Virginia (37th St. WTP) (CDM,

1989), Wichita, Kansas (lime sludge), North Marin County

Water District, and Novato, California (U.S. EPA,

1978a).

6.3.2.5 Truck Hauling

A primary consideration in determining whether to use

truck hauling to transport liquid WTP residuals is the

potential impact, perceived or real, of residuals on the

WWTP liquid stream processes. When a significant impact is anticipated, the use of truck transport allows the

separate introduction of WTP residuals into a process

or receiving vessel at the WWTP, where negative impacts can be minimized. Distance to the WWTP and

quantity of residuals, as well as the congestion of land

use, often determine whether to use a pumping/force

main system or truck haul. If residuals quantities are

small or if distance to the WWTP exceeds 10 miles,

truck haul is preferred over pumping through a force

main. Additionally, if the WTP residuals have already

been thickened or dewatered, truck transportation is

often the only feasible method of delivering the residuals

to the WWTP.

Trucking partially dewatered WTP residuals often results in solids compression because of vibration and

release of free water (ASCE/AWWA, 1990). As a result,

the truck bed should be sealed and watertight. Wastewater treatment facilities are well aware of the requirements for truck transport of sludges from their facilities,

and the same precautions and limitations should be

applied to truck transport of WTP residuals, whether

dewatered, thickened, or unthickened.

6.4 Design Considerations for

Wastewater Treatment Plants

Anything discharged to a WTP, including WTP residuals,

has an impact on the design and operation of the facility.

Careful consideration of the design and/or operation of

the WWTP, however, can mitigate the effects of discharging WTP residuals to the facility.

The first issues to consider are the types and characteristics of WTP residuals to be discharged to the

WWTP. These include not only the physical and chemical characteristics of the residuals, but the form in which

they are to be conveyed to the WWTP (i.e., as a liquid,

semi-solid, solid.

The next important factor to consider is the manner in

which the residuals are to be conveyed to the WWTP.

Typical conveyance systems include gravity sewers,

pumping/force mains, and truck hauling. The type of

conveyance system used is important because it determines where the WTP residuals can be introduced into

the WWTP’s processes (i.e., to the liquid and/or solids

handling processes), as well as the rate at which the

residuals are to be introduced (i.e., gradually over a long

period of time, or as a slug dose).

The points at which WTP residuals are introduced to the

WWTP for processing must be carefully considered. In

109
certain cases, introducing residuals to the liquid treatment processes may have an adverse effect on those

processes while, in other situations, it may improve the

performance of the processes. Sometimes, it may be

more advantageous to introduce the residuals to the

WWTP’s solids handling processes to prevent overloading the liquid treatment processes. Introducing WTP

residuals to the solids handling processes does not

necessarily eliminate all impacts on the liquid treatment

processes because this option can still affect the recycled flows to the liquid treatment processes.

Obviously, the final and most important consideration in

accepting WTP residuals at a WWTP is the impact on

the performance of the WWTP. The performance of the

unit treatment processes at the WWTP must be maintained to ensure that discharge of the final effluent to the

receiving body is not adversely affected, and that disposal or beneficial reuse of the residuals from the

WWTP does not become a problem.

As noted in Chapter 2, the NPDES permit governs any

direct discharges to waters of the United States. The

introduction of WTP residuals to a WWTP may affect two

areas of the WWTP’s NPDES permit: the parameters for

which specific effluent limits are described, and the

Whole Effluent Toxicity (WET).

Equally important to meeting effluent NPDES permit

requirements is the fact that the disposal and/or reuse

of residuals from a WWTP has taken on new meaning

as a result of the promulgation of the 40 CFR Part 503

sewage sludge regulations (U.S. EPA, 1993c). These

regulations address the disposal and beneficial reuse of

sewage sludge in three general categories—land application, surface disposal, and incineration.

Constituents in WTP residuals discharged to a WWTP

may affect the quality of the WWTP residuals as governed by 40 CFR Part 503, especially the concentrations of 10 heavy metals. If aluminum is included in

round 2 of the 40 CFR Part 503 regulations, it definitely

will present a problem for discharge of WTP residuals to

WWTPs. The remaining sections of this chapter discuss

design issues that must be carefully considered when

evaluating whether to accept WTP residuals in the liquid

or solids handling processes at a WWTP.

6.4.1 Hydraulic Loading

Hydraulic loading is generally not a factor in the acceptance of WTP residuals conveyed by gravity sewers to

a WWTP. Large volumes of liquid wastes generated by

WTP processes in a short period (e.g., filter backwash

waters, IX regenerate waste, reject water from membrane processes), are usually equalized at the WTP

prior to being discharged to a sewer. This is done to

prevent overloading of the gravity sewer itself. If large

volumes of liquid WTP residuals are not equalized before discharge, attenuation that takes place in the sewer

generally prevents a hydraulic loading problem at the

WWTP.

It is generally more important to check the hydraulic

loading of WTP residuals conveyed to a WWTP through

a dedicated pump/force main or by truck. The hydraulic

loading on unit treatment processes at the WWTP is

usually not the controlling parameter when WTP residuals are introduced to the liquid treatment process train,

unless the WWTP is a small one.

The condition that most warrants concern about hydraulic loadings is when WTP residuals are discharged to the

solids handling processes at a WWTP. In this case,

equalization at either the WTP or WWTP is generally

required.

6.4.2 Organic Loading

The organic content of WTP residuals varies widely,

depending heavily on the quality of the raw water processed at the WTP (Albrecht, 1972; Calkins and Novak,

1973; Cornwell et al., 1987). The concentration of organic matter in WTP residuals, as measured by BOD or

COD, is not typically in the range of that found in wastewater. Although organics loading is typically the controlling parameter for the biological unit treatment processes

in both the liquid and solids handling process trains at a

WWTP, the additional organics loading encountered by

introducing WTP residuals generally does not have a

significant impact.

6.4.3 Solids Loading

The impact of solids in WTP residuals on the WWTP is

limited to two general areas:

• The performance of primary, intermediate, and final

clarifiers.

• The constituents in the WWTP biosolids as related

to the requirements of the 40 CFR Part 503 sewage

sludge regulations.

The introduction of additional solids from WTP residuals

generally does not significantly affect the ability of the

WWTP to comply with its NPDES permit requirements

for the final effluent (McTigue et al., 1989; Tsang and

Hurdle, 1991). In several cases, alum and iron residuals

from WTPs have actually improved the efficiency of

primary clarification in WWTPs; in another case, these

residuals increased the amount of phosphorus chemically precipitated from a WWTP (McTigue et al., 1989;

Tsang and Hurdle, 1991). A common factor related to the

discharge of WTP residuals to a WWTP is an increase

in biosolids from the WWTP, with a corresponding decrease in the volatility of the biosolids.

In certain instances, metals present in WTP residuals

can affect NPDES permit effluent limits. This situation is

110
becoming more of a problem since most of the specific

effluent requirements for metals in NPDES permits have

daily limits. If a significant portion of the precipitates in

the WTP residuals is in colloidal form, the solids pass

through primary and/or secondary clarification at the

WWTP.

Another cause of increased metals concentrations in a

WWTP effluent can be the resolubilization of metals that

occurs when WTP residuals are processed in a WWTP’s

solids handling processes, and recycled flows from solids handling are returned to the liquid process train.

Precipitates and other solids in WTP residuals are fairly

inert. Concern has been expressed that additional solids

from WTP residuals could impair fixed film biological

treatment processes, but definitive information to substantiate this concern is unavailable.

6.4.4 Toxics Loading

The potential toxicity of WTP residuals is a concern

because of 1) the potential inhibitory impact on biological unit treatment processes at the WWTP, and 2) the

possibility of failing a bioassay with Ceriodaphnia dubia

or fathead minnows, resulting in noncompliance with the

WET requirements.

A discussion of the potential for hazardous waste classification of WTP residuals is presented in Chapter 2.

Passing the TCLP test allows a WTP residual to be

classified as nonhazardous; however, it does not preclude the possibility of toxic effects on a WWTP if the

WTP residual is discharged to the facility. Although organics and inorganics can be the cause of toxic effects

from WTP residuals, heavy metals are most often responsible for toxicity problems at WWTPs.

Nearly all heavy metals can exert toxic effects at elevated concentrations. Heavy metals can be grouped into

“essential” and “nonessential” metals, according to their

importance in biological systems. The vast majority of

heavy metals fall into the “essential” category. Cadmium, mercury, and lead are the most commonly found

nonessential heavy metals. In assessing the potential

toxicity of individual heavy metals, two important issues

should be considered:

• For metals essential to biological processes, an increase in concentration can often improve biological

conditions if a deficiency of metals existed initially.

• When they are present in excess, the essential metals can sometimes exert greater toxicity than the nonessential heavy metals.

Metal speciation plays a major role in the potential toxicity of a particular heavy metal. The dissolved portion,

available as the free metal ion, is generally the most

toxic to biological systems. This simple differentiation

between the dissolved and particulate forms, however,

is generally not adequate to determine toxicity effects.

Heavy metals in the dissolved phase can be free metal

ions, or can be tied up with inorganic or organic compounds to form complexes. Heavy metals in the particulate phase are capable of being bound to the surface of

other solids, or tied up in the bulk phase of a precipitate.

6.4.5 Liquid/Solids Separation

A key impact of WTP residuals on WWTP unit processes

is that additional residuals handling will be necessary.

The effects of the additional residuals on clarification,

digestion, dewatering, and final disposal must be considered. Generally, final biosolids volume increases as

the amount of influent WTP residuals increases. An

increase in the primary clarifier sludge volume is to be

expected with the addition of WTP residuals to raw

wastewater.

Liquid/solid separation is an important unit operation in

water and wastewater treatment. The success of the

overall treatment process depends greatly on a facility’s

ability to separate solids from the liquid stream. Liquid/

solid separation can be accomplished by several unit

operations. This section discusses the primary and secondary sedimentation operations, while other liquids/

solids operations are discussed in Section 6.4.6.3.

When WTP residuals are discharged into a sanitary

sewer system, one of the first unit processes that they

encounter is primary sedimentation at a WWTP. Several

studies conducted on the impact of WTP residuals on

the sedimentation process have reported fairly consistent results.

Full-scale testing at the Hookers Point Sewage Treatment Plant in Tampa, Florida (Wilson et al., 1975)

showed that when WTP residuals were discharged to a

sanitary sewer system, a fluffy blanket layer of sludge

formed at a level somewhat above the normal blanket

level in WWTP primary sedimentation tanks. At times,

the sludge collectors in the primary sedimentation tanks

seemed incapable of collecting this fluffy material even

when the sludge pumps were run at their maximum rate.

Data show that at a WTP residuals dosage of up to 50

mg/L, removal of suspended solids from the primary

sedimentation tanks improved. At dosages above 50

mg/L, the percentage removals may be expected to

drop. Wilson et al. (1975) note that when WTP residuals

were accepted on a regular basis at the plant, the primary sludge solids content dropped from 5.8 to 4.7

percent. The addition of high concentrations of alum

residuals tended to produce a primary sludge that was

difficult to handle. Dosages of fresh alum less than 100

mg/L or WTP residuals of 40 mg/L were reported to

produce a more manageable product.

Rolan and Brown (1973) report that combined alum

residuals and sewage sludge did not settle as well as

111
each would separately. Sludge volume tests were conducted for different dosages of WTP alum residuals to a

WWTP (see Figure 6-1). The addition of 200 mg/L of

alum residuals approximately doubled the volume of

primary clarifier sludge. It was concluded that WTP residuals in concentrations of 200 mg/L or less would have

no significant adverse effects on a WWTP.

Rindt (1973) studied the effects of alum on primary

sedimentation in a series of laboratory tests using laboratory-prepared alum and WTP alum residuals. He reports that WTP alum residuals caused nearly a four-fold

increase in the sludge volume index (SVI) when added

to wastewater samples. Turbidity and total suspended

solids in the supernatant also increased in proportion to

the amount of residuals disposed of and the suspended

solids concentration of the residuals.

Hsu and Pipes (1973) investigated the effect of aluminum hydroxide floc on the settling characteristics of

primary and secondary effluent. Settleability of bulking

sludge, as measured by the SVI, may be improved

significantly with the addition of aluminum hydroxide, but

there will be no effect on the SVI of normal sludge. The

authors attribute the decrease in the SVI of bulking

sludge to the inorganic aluminum hydroxide enmeshment, which increased the density and the compatibility

of the sludge.

Nelson et al. (U.S. EPA, 1978a) conducted a full-scale

evaluation of alum residuals discharged via sanitary

sewer to an activated sludge WWTP and found that the

efficiency of primary settling decreased by approximately 10 percent. Phosphorus removal improved by

approximately 12 percent, and scum removal and settling improved in secondary clarification.

Camp Dresser & McKee, Inc. (CDM, 1989), conducted

a study of Norfolk, Virginia’s, 37th Street Water Treatment Plant residuals and their subsequent treatment at

the Hampton Roads Sanitation District (HRSD) Virginia

Initiative Plant (VIP). To verify the potential effect that

WTP residuals might have on the VIP, representative

quantities of the residuals were hauled by truck to two

of HRSD’s smaller plants at a rate comparable to VIP

receiving all of the residuals. The sludge was processed

through liquids and solids processing facilities at one

plant, and through liquid stream facilities at the other.

The study showed the following:

• The addition of WTP residuals to the WWTP influent

caused the primary sludge concentration at the

WWTP to drop from 3.5 to 2.5 percent, increasing

the volume of liquid primary sludge to be pumped by

approximately 60 percent.

• The addition of WTP residuals to the WWTP influent

decreased the efficiency of solids settling in the secondary clarifiers, resulting in the need for additional

clarifier capacity or the addition of a polymer to enhance settling.

• Discharging of WTP residuals directly to the WWTP

solids handling facilities resulted in minimal adverse

impacts on solids processing facilities other than increased flows.

The study concluded that sanitary sewer discharge of

WTP residuals to the VIP was not feasible because of

the adverse effect on settling at VIP and the need to

obtain an exception to the HRSD Industrial Discharge

Regulations as a result. The processing of WTP residuals, transported by force main to the VIP and introduced

directly to the solids processing facilities (i.e., centrifuges), appears technically feasible. The residuals will

be dewatered separately by the centrifuges and then

combined with the WWTP biosolids before being fed to

the incinerators.

6.4.5.1 Summary

All the above studies involved WTP coagulation residuals, which constitute most of the WTP residuals produced in the United States. The findings from these

studies confirm that 1) the WWTP sludge volume will

generally increase, 2) phosphorus removal from the

wastewater will be improved, and 3) sludge concentration in the underflow may decrease when coagulation

sludge is discharged to a WWTP. Lime softening residuals settle very well and should not affect the wastewater

treatment primary settled sludge volume as much as

coagulation residuals do.

Figure 6-1. Effect of WTP sludge on the combined volume of

wastewater sludge after 30 minutes of settling 

(Rolan and Brown, 1973).

112
When evaluating the option of discharging WTP residuals to a WWTP via a sanitary sewer system, sedimentation tank capacities must be checked to ensure that

they will not be overloaded by the additional solids. This

precaution is especially critical for primary sedimentation, since the majority of the residuals solids will be

removed here. The capacity of sludge pumps, pipes,

and subsequent conveyance systems must also be examined; the removal of a larger volume of sludge at a

lower solids concentration would require a larger pumping capacity and/or a higher pumping frequency.

6.4.6 In-Plant Solids Handling

The introduction of WTP residuals to a WWTP will inevitably increase the solids loading to the whole treatment

train. This is especially critical for the solids handling

operations and treatment processes such as sludge

digestion, thickening, and dewatering.

6.4.6.1 Aerobic Digestion

Aerobic digestion is a sludge stabilization process used

primarily in small treatment facilities. The process is

similar to the activated sludge process. Sludge in the

digester is aerated to oxidize the organic substrate, and

as the available substrate is depleted, the microorganisms begin to consume their own protoplasm through

endogenous respiration.

Important design criteria for aerobic digesters include

hydraulic detention time, solids loading, oxygen requirement, and energy requirements for mixing. Both volatile

solids loading together with total solids loading should

be considered. Because aerobic digestion is similar to

the activated sludge process, the effect of WTP residuals on aerobic digestion should be somewhat similar to

the impacts of the residuals on activated sludge systems. More data are available on the impact of WTP

residuals on activated sludge, than on aerobic digestion.

Consequently, some of the following discussion will be

on activated sludge systems.

To accommodate WTP residuals, an aerobic digestion

system must have the physical capacity to accept the

additional solids loading that residuals bring. Although

the volatile solids contribution from the residuals is expected to be low and should not significantly affect the

overall volatile solids loading to the digester, the impact

of the residuals on hydraulic and solids retention time,

as well as the overall reduction in volatile solids concentration, must be considered.

Toxic compounds present in WTP residuals can adversely affect the biological processes of wastewater

treatment facilities. Dissolved solids present in a liquid

phase waste may thwart the biological process, depending on their form and/or concentration. Predischarge

equalization techniques may alleviate toxicity problems,

as a continuous discharge of suitable dilution will enhance the ability of the microorganisms present in the

biological process to adapt and adjust to the presence

of any inorganic ion. If the dosing of WTP residuals is

equalized so that surges do not occur, and the dose is

kept below 150 to 200 mg/L, no direct effect on the

activated sludge process is likely to take place, although

downstream process effects or solids handling process

effects are possible (van Nieuwenhuyze et al., 1990).

Several investigations have found no evidence that

alum hinders the substrate removal properties of the

activated sludge process. Nelson et al. (U.S. EPA,

1978a) reported no change in the efficiencies of COD

and BOD removals when alum sludge was discharged

into a wastewater treatment plant. In a series of laboratory-scale studies, Voorhees (1974) studied the effects

of alum sludge on the activated sludge process and

reported that no significant effects on either COD or

TOC removal were observed at alum dosage of up to

100 mg/L. Voorhees observed that aluminum hydroxide

sludge did not decrease SVI, and that the settling characteristics were not adversely affected in the secondary

clarifier. When a shock load of alum sludge was introduced to the system, the COD and TOC removals were

decreased. The system was able, however, to recover

to the same COD and TOC removal efficiencies before

the shock load.

Despite several findings reporting the apparent lack of

impact of WTP residuals on biological treatment processes, the metal contents of these residuals do pose

some concerns. Reid et al. (1968) presented data to

show that metallic ions such as hexavalent chromium,

cadmium, copper, nickel, aluminum, and silver can hinder biological processes.

Hsu and Pipes (1973) reported that doses of aluminum

hydroxide in the range of 10 to 300 mg/L as aluminum

to an aeration tank caused a significant increase in total

phosphorus removal, no change on the nitrification, and

a slight increase in COD removal. Hsu and Pipes stated

that the most significant adverse effect on the biological

treatment process was the increased volume of sludge

produced. As a result, the sludge withdrawal rate and

sludge return rate would have to be controlled very

carefully to ensure adequate supply of active biological

sludge mass to the aeration tank.

Anderson and Hammer (1973) studied the effects of

alum addition on activated sludge biota and found that

the bacteria were able to synthesize soluble organics in

a chemical-biological sludge process and that BOD removal was not affected. Higher life forms such as protozoa, however, found the environment under high alum

additions too adverse for normal existence.

The level of salt concentration in WTP residuals is extremely important to the proper performance of the bio113
logical processes at a WWTP. Introducing WTP residuals with high sodium chloride concentrations, such as

may occur from the undiluted discharge of IX waste,

could cause considerable stress to biological organisms. As a result, COD of the biological reactor effluent

could increase after the salt addition. Threshold concentrations of toxic ions in the biological process are the

level above which a decrease in the COD removal efficiency of the biological process occurs. If precautions

for equalization and dilution of WTP residuals prior to

discharge are adhered to, then brine discharges are

unlikely to affect the receiving wastewater plant detrimentally (Cornwell et al., 1987).

6.4.6.2 Anaerobic Digestion

Anaerobic digestion is one of the oldest processes used

to stabilize sludges. It involves the decomposition of

organic and inorganic matter in the absence of oxygen.

Important design criteria for anaerobic digestion include

hydraulic and solids retention times, volatile solids loading, mixing energy, gas production, and volatile solids

reduction.

The additional volume and dry weight of sludge and the

decrease in percent of volatile matter must be considered when determining the sizing and performance parameters of digesters. A 1989 study (Foley et al., 1989)

of the East Bay Municipal Utility District (EBMUD) in San

Francisco, California, examined the effects of alum residuals on a digester by placing the alum residuals

directly into the digester. The performance of digester 11

receiving alum residuals directly was compared with

digesters operating under “normal” procedures (i.e.,

treating a mixture of primary and WTP residuals introduced into the influent).

The results of this part of the study are presented in

Table 6-3. The alum residuals that were fed directly to

digester 11 had an average total solids (TS) percent of

2.5 and an average volatile solids of 1 percent TS. No

data were presented on the rate of gas production versus volatile solids reduction. There was a significant

increase in aluminum concentration in the digester receiving the alum residuals compared with the digesters

operating under “normal” procedures. This needs to be

considered relative to ultimate disposal options. All other

parameters were essentially equal.

Because softening residuals can cause lime deposition,

discharging softening residuals to a WWTP may adversely affect the performance of the anaerobic digester.

If softening residuals are added directly into the anaerobic digester, the temperature can fall because of the

additional volume of inert material. Additional heating

may be required in this event.

A study in Durham, North Carolina (McTigue et al., 1989)

investigated the impact of controlled alum residuals discharge to a sanitary sewer on the physical facilities at a

WWTP. The study showed an insignificant effect on flow

rate through the wastewater treatment system. The digested solids, however, increased 46 percent with the

addition of WTP residuals. The loadings on the biosolids

drying beds increased by the same percentage, from 9.8

lb/ft2

/yr to 14.3 lb/ft2/yr, with the addition of these residuals. The retention time and surface loading rate associated with other processes were not significantly

affected, with the exception of the volume of sludge

collected and handled in the primary clarifier, which was

estimated to increase by 59 percent, from approximately

3.91 to 6.2 tons per day (TPD). The retention time in the

aerobic digesters decreased from 34.7 to 26.7 days, a

23 percent reduction, when WTP residuals were introduced. The impact on the wastewater treatment facility

was significant, particularly in the retention time of the

aerobic digesters, increased pumping needed for the

primary clarifier sludge, and the additional drying bed

capacity required.

Rindt (1973) used a laboratory-scale study to evaluate

the impact of alum residuals on anaerobic digestion. He

found that gas production and TOC remained largely

unaffected. Total solids in the digester increased because of the additional alum residuals load. The pH in

the digester was also depressed slightly by the alum

residuals. The percentage of volatile solids in the digester decreased because of the increased inorganic

input. Inorganic phosphates in the digester were almost

entirely precipitated by the aluminum present.

These results were confirmed by Reed (1975), who

used laboratory-scale anaerobic digesters to study the

Table 6-3. Comparison of Digester 11 With Background

Digesters (Foley et al., 1989)

Parameter

Digester 11

(mg/L)

Background 

Digesters

(mg/L)

Aluminum 15,000 7,200

Cadmium 16 18

Chromium 140 140

Copper 390 400

Iron 16,000 17,000

Lead 180 160

Nickel 41 41

Silver 16 23

Zinc 900 960

pH 7.7 7.6

TS 2.3 2.5

VS 1.4 1.6

TS/VS 58 61

114
effects of both reagent grade alum (aluminum sulfate)

and WTP residuals (aluminum hydroxide) on the digestion process. Under comparable aluminum concentrations, Reed found that digestion was adversely affected

in the digester where feed sludge contained aluminum

sulfate, but not in the digester where feed sludge contained aluminum hydroxide residuals. Inhibition was attributed to the high sulfide concentration associated with

the reagent grade alum.

Barth et al. (1965) studied the effect of metals on biological processes and found that for chromium, nickel,

and zinc, an influent metals concentration of 10 mg/L—

for any one or combination of the three—did not affect

digestion. Copper continuously present at 10 mg/L

caused failure of combined sludge digestion.

Wilson et al. (1975) did not observe any negative impacts at the anaerobic digesters in Tampa, Florida, except that the solids content of the biosolids discharged

to the drying beds was reduced by a factor of approximately two as a result of the WTP residuals addition.

Nelson et al. (U.S. EPA, 1978a) observed an increase

in digester gas production beyond that generated by the

increased load of volatile solids applied. Nelson et al.

speculated that this was due to easier mixing of the

digester contents as a result of WTP residuals being

added. Hsu and Pipes (1973) reported that aluminum

hydroxide retarded the digestion process.

Emig (1979) conducted a full-scale test employing alum

and polymer to treat phosphorus at a WWTP. He reported a drop in digester gas production, accompanied

by a drop in digester pH (from approximately 7.2 to 6.0)

after the addition of alum. This change led to a digester

upset at alum dosage of approximately 400 mg/L. Alum

addition had reduced the alkalinity of the digesting

sludge and increased the volatile acid levels. The upset

was caused by the depression of the alkaline wastewater pH.

These studies show that unless alum residuals are introduced at a very high rate, anaerobic digesters are

unlikely to be significantly affected, provided the digester has adequate capacity to accommodate the increased solids loading. In assessing digester capacity,

the reduced volatile solids concentration of the residuals

stream must be taken into account.

6.4.6.3 Thickening/Dewatering

Thickening and dewatering are unit operations that

separate liquid from solids. In its unthickened condition

(e.g., underflow withdrawal from a clarifier), a WTP residuals/WWTP sludge blend normally has a solids concentration in the range of 2 to 6 percent total solids.

Thickening can reduce the water content to the range of

9 to 15 percent total solids. In this range, the thickened

residual still behaves like a liquid but has the consistency and physical properties of chocolate pudding. Dewatering of thickened residuals will result in a cake with

a solids content of 18 to 40 percent.

Thickening and dewatering can be accomplished by

natural or mechanical means. Examples of thickening

operations are gravity thickening, flotation thickening,

centrifuge thickening, and rotating drum thickening.

Common dewatering operations include sand or paved

drying beds, dewatering centrifuges, belt filter presses,

and recessed chamber filter presses. All thickening and

dewatering equipment is sized according to solids loading. The introduction of WTP residuals may cause overloading of these facilities, depending on the quantity of

residuals introduced and the existing WWTP loadings

relative to the plant’s design capacity.

Hsu and Pipes (1973) reported that the dewaterability of

wastewater biosolids, as measured by the specific resistance to filtration, was improved by an increase in the

dosage of aluminum hydroxide. Nelson et al. (U.S. EPA,

1978a) found that the concentrate quality and cake solids content of the centrifuge were not significantly

changed by the introduction of WTP residuals. They

noted that the yield from the centrifuge increased from

600 to 800 lb/hr for biosolids, to about 900 lb/hr for the

alum residuals biosolids mixture. A study in Norfolk,

Virginia (CDM, 1989), showed that the settling characteristics of waste activated sludge as measured by the

sludge volume index was negatively affected by the

addition of alum residuals. No significant impacts on

centrifuge dewatering in terms of cake solids, polymer

dosage, and centrate recovery were reported.

In almost every case, the dewaterability of the wastewater biosolids, as measured by the specific filtration resistance, was shown to improve with the addition of WTP

residuals (van Nieuwenhuyze et al., 1990). A study by

McTigue et al. (1989) investigated how the use of WTP

residuals changed sludge handling characteristics and

affected WWTP performance. A pilot column evaluation

was conducted to determine the effect of WTP residuals

on drainage characteristics of the WWTP biosolids.

These tests were necessary to determine the impact of

the WTP residuals on the WWTP’s ultimate dewatering

process (i.e., sand drying beds). McTigue et al. (1989)

found that the addition of WTP residuals had little influence on the drainage characteristics of the aerobically

digested activated sludge. The addition, however, of

residuals to aerobically digested activated sludge

caused the biosolids to settle in the bed and produce a

clear water layer suitable for decant, thus improving the

dewatering process. The dewatering tests on the combined anaerobically digested primary sludge and WTP

residuals were very similar. The major difference was

that the anaerobically digested primary sludge drained

faster than the aerobically digested activated sludge and

115
also contained some clarified surface water available for

decant.

In Allentown, Pennsylvania, a pilot test was conducted

to determine the effects of dewatering alum residuals

either separately or combined with biosolids using a

recessed chamber filter press (Koplish and Watson,

1991). The dewatering results indicated that the highest

cake solids of 46.8 percent were achieved by dewatering alum residuals alone, although a cycle time of 3

hours was required. In comparison, wastewater

biosolids were dewatered to 34.3 percent cake at a cycle

time of 1 hour. When they were mixed at different proportions, the resultant cake solids were between 34.3 to

46.8 percent; the blend having higher portions of alum

residuals resulted in the higher cake solids.

With the success of the pilot tests, the City of Allentown

has proceeded to install full-scale facilities for accepting

and dewatering a combination of biosolids and WTP

residuals at its WWTP. The operation is termed the

“Centralized Dewatering and Co-Disposal (CDCD)” concept, and is described by Koplish et al. (1995). Dewatering of a 50-50 blend is accomplished at the WWTP

by high-solids belt filter presses, and the combined operation is saving the city at least a quarter of a million

dollars annually. At a 50-50 blend, feed solids in January

and February 1995 were 2.6 percent; cake solids were

25.6 percent. The WTP residuals are tank-trucked to the

WWTP, where they are discharged to a blended residuals holding tank and mixed with biosolids. The operation

has been so successful that the city is now considering

accepting third party liquid residuals for processing,

which could represent a new revenue source for the city.

A pilot study was conducted in Wilmington, North Carolina, employing a belt filter press.4

 It was found that alum

residuals dewatered alone attained a maximum cake

solids content of 23 to 27 percent. Wastewater biosolids

were dewatered to 15 to 18 percent solids. When alum

residuals and wastewater biosolids were blended, the

dewatered cake had a solids concentration of 17 to 20

percent. The results were consistent with bench-scale

tests conducted earlier on the same sludges. The reported test results followed a pattern similar to the Allentown study.

Vandermeyden and Potter (1991) reported on full-scale

sludge dewatering tests for alum WTP residuals using

both centrifuge and belt filter press equipment. The tests

yielded favorable results in terms of cake solids concentration, polymer dose, and centrate/filtrate quality. The

dewatering test results are summarized in Table 6-4 and

indicate that cake solids concentrations typically ranged

from 20 to 27 percent, polymer dose from 5 to 70 lb/ton

of solids, and overall solids capture around 98 percent.

Although available information has generally shown that

WWTP biosolids dewaterability is not negatively impacted by alum residuals, gravity settling of the wastewater sludge may be affected. The ability to reduce the

water content of sludge depends on many factors, including the physical, chemical, and biological characteristics of the sludge, and the unit operations employed

to thicken or dewater it. To date, the science of dewatering has not advanced to the point when sludge dewaterability can be predicted precisely without actual testing.

Data that are useful in determining sludge dewaterability, such as particle size distribution and zeta potential,

are seldom available.

If the impacts of cake solids are critical to downstream

processing, the effect of WTP residuals on sludge thickening and dewatering should be fully investigated

through pilot- or full-scale testing before the co-processing of WTP residuals and WWTP biosolids is implemented.

6.5 Ultimate Disposal of Wastewater

Treatment Plant Biosolids

Several different ultimate disposal options are available

to WWTPs. Before the advent of the Clean Water Act

(CWA), one common disposal option was direct discharge to a nearby surface water body. Current regulations limit this as a viable option for WWTPs. The

regulatory climate further requires WWTPs to develop,

evaluate, and implement alternative ultimate disposal

methods. Options include land application, incineration,

and composting, all of which could be considered beneficial uses. The regulations governing beneficial reuse

options have only recently been established by some

primacy agencies; other primacy agencies are still developing regulations. The promulgation of 40 CFR Part

503 will further these options as guidelines and regulations are established.

Quality, quantity, and cost are key factors in determining

which disposal options a WWTP employs. The regulatory requirements for WWTP biosolids, established by

the primary agencies responsible for monitoring ultimate

disposal, are applied to WTPs that discharge to

WWTPs. Regulatory requirements and technical issues 4 City of Wilmington. 1991. Personal communication.

Table 6-4. Alum Treatment Plant Sludge Dewatering Test

Results (Vandermeyden and Potter, 1991)

Parameter

Belt Filter

Press Centrifuge

Loading rate (lb/hr) 700–800 560–750

Feed solids (%) 2.5 2.5

Polymer dose (lb/hr) 20–23 25–27

Solids capture (%) 98 98

116
relating to these disposal options are addressed in the

following sections.

6.5.1 Direct Discharge

In the past, direct discharge was commonly used to

dispose of WTP residuals. Protection of aquatic resources and maintenance or improvement of water

quality conditions has resulted recently in more stringent

regulation. As a result of the CWA, regulation of this

option has increased the cost and concern over liability

to the WTP. The regulatory requirements are primarily

associated with the Federal Ambient Water Quality Criteria, established through the CWA to protect aquatic life

in marine and freshwater environments. Regulations of

particular interest to the water treatment industry are

those that establish acceptable levels of metals, suspended solids, and water treatment–related contaminants (see Chapters 2 and 5).

For a WWTP to consider receiving WTP residuals, anticipated levels of metals and other contaminants would

have to be determined to ensure compliance with

NPDES permits limits.

6.5.2 Land Application

Land application, as it applies to disposal of WTP residuals, is discussed in Chapter 8. The same quality restrictions and load limitations that apply to land application

WTP residuals apply to wastewater biosolids; however,

biosolids have the added benefit of higher organic or

nutrient content. Higher organic content, nitrogen, and

phosphorus levels are beneficial if land application of

the biosolids is associated with agricultural uses.

The addition of alum WTP residuals to WWTP biosolids

could reduce the level of available phosphorus and

create concerns over aluminum toxicity. The addition of

residuals from a water softening WTP, or the addition of

alum residuals conditioned by lime for dewatering could

prove to be very beneficial to a joint land application

program with biosolids. Examples of each of these two

approaches are reported by Parsons and Waldrip (1995)

and by Assadian and Fenn (1995).

Ionic aluminum is toxic to many plants, including agricultural crops (Elliot and Singer, 1988). This detrimental

effect could limit the land application process for

biosolids containing alum WTP residuals. A WWTP

manager currently land applying biosolids must therefore be watchful of additional aluminum in the biosolids

to ensure that regulatory limits are not exceeded. Recent studies, however, performed by Grabarek and Krug

(1989) show that large doses of alum sludge do not

necessarily produce aluminum toxicity (Elliot et al., 1988).

As discussed previously, one benefit of land applying

WWTP biosolids is the contribution of organic matter

and nutrients. The addition of aluminum or iron hydroxides could result in a strong fixation of available phosphorus (PO4

-3), thus making it unavailable for vegetation.

This does not preclude the addition of WTP residuals,

but careful consideration must be given to relative quantities and qualities of any additives, as well as beneficial

reuse objectives. If land spreading for nonagricultural

purposes is the ultimate disposal method, concerns over

aluminum toxicity and available phosphorus are minimized. Pilot-scale testing and chemical analysis can be

conducted to assess viability.

The City of Boulder, Colorado, has undertaken an extensive evaluation of the benefits and risks of coapplication of WTP residuals and WWTP biosolids (Harberg

and Heppler, 1993). They reported that the beneficial

effects of direct addition of WTP residuals to WWTP

biosolids depend on soil conditions, characteristics of

the residuals, and application rates. If not properly managed, direct application of WTP residuals could cause

detrimental plant growth and other environmental effects. An evaluation of the literature on the effects of

WTP residuals on crop growth response suggested that

land coapplication of WTP residuals and WWTP

biosolids for beneficial use was a viable way to recycle

WTP residuals, and did not reduce the plant-available

phosphorus in soils. In fact, the WWTP biosolids would

help supplement the phosphorus available in the soil. A

proper combination of the two materials could also provide the nitrogen needed by the plants without reducing

the phosphorus availability in the soil. The City of Boulder study is specifically evaluating coapplication of an

aluminum-based WTP residual and a WWTP biosolid to

dry land winter wheat. Based on crop yield, the WTP

residuals did not significantly affect yields, and no difference between control and applied plots was found.

6.5.3 Incineration

Incineration provides an effective method for reducing

volume and creating a relatively inert material. Depending on the initial metals levels in the biosolids, leaching

of metals from the ash residue may be a concern.

The trend in biosolids disposal is increasingly in favor of

beneficial reuse options. Incineration is becoming a less

desirable option because of its relatively higher costs, as

well as recent concerns and regulations on air emissions.

The addition of a WTP residual into the processes of a

WWTP currently incinerating biosolids could create

some problems. Because the organic content of WTP

residuals is generally lower than that of WWTP

biosolids, the thermal destruction process may be hindered. The presence and contribution of iron and aluminum, as well as other coagulant contaminants, must

also be considered relative to facility operations, air

emissions, and leaching from ash residue.

117
6.5.4 Composting

Composting is a biological process that converts

biosolids into a stable humic material that could be land

applied (Tsang and Hurdle, 1991). Composting (including co-composting) of WWTP biosolids, considered a

beneficial reuse option, has been increasing as a disposal practice in the United States (Bowen et al., 1991).

The problems associated with composting WWTP

biosolids are generally related to the presence of pathogens and odors. The addition of WTP residuals to the

WWTP waste stream would not compound these problems, and could act to dilute them. The resulting lower

volatile solids content of the biosolids, however, could

inhibit the degradation process (Tsang and Hurdle,

1991) by reducing the level of biological activity.

Studies conducted in Greenwich, Connecticut, and Myrtle Beach, South Carolina, indicate that if managed properly, composting WTP residuals and WWTP biosolids

could be an attractive alternative.

Vandermeyden and Potter (1991) reported on full-scale

tests of joint composting of alum residuals and wastewater biosolids. A 25-percent mix ratio (residuals/

biosolids) appeared to be the maximum ratio for attaining proper composting temperatures for pathogen destruction and for maintaining levels of volatile solids. No

observable differences in color or physical characteristics were noted between compost with mix ratios of

25 and 12.5 percent and virgin biosolids compost. The

introduction of metals from the alum residuals did not

produce a degradation in compost quality and actually

caused a higher allowable loading rate for the finished

compost, based on a lowering of the copper and zinc

concentrations in the final compost.

118
Chapter 7

Landfill Options

As a drinking water utility evaluates its treatment plant

residuals management options, landfilling will most

likely be considered. In nearly every case, landfilling

falls into two categories: co-disposal and monofilling.

This discussion assumes that the water treatment plant

(WTP) residuals to be managed have not been classified as hazardous wastes. When the WTP residuals

exhibit hazardous characteristics, they must be handled

as RCRA Subtitle C wastes. In most instances, WTP

residuals will be classified as nonhazardous solid waste,

subject to the requirements of RCRA Subtitle D.

Under RCRA Subtitle D regulations (40 CFR Parts 257

and 258), criteria have been established for the design

and operation of nonhazardous solid waste landfills. The

requirements of 40 CFR Part 257 apply to landfills that

receive only WTP residuals (monofills), as well as landfills that accept solid waste other than household waste

(e.g., industrial waste). These criteria, which are performance based and do not include any specific design

criteria, address seven landfill design and operation areas (see Table 7-1).

Municipal solid waste landfills (MSWLFs) are subject to

the criteria of Part 258. Like the provisions of Part 257,

Part 258 covers performance-based criteria, but it also

includes specific design criteria. If a utility disposes of

its drinking water residuals in a monofill, then Part 258

criteria do not apply. If, however, the WTP residuals are

co-disposed of with municipal solid waste, including

household waste, the requirements established for

MSWLFs apply. The landfill criteria in 40 CFR Part 258

address six major areas, as listed in Table 7-2.

RCRA Subtitle D establishes minimum criteria to be

applied in all states. In addition, each state may develop

its own landfill requirements provided that the criteria at

least meet the minimum federal standards. In many

states, landfill requirements are more restrictive than

federal criteria. Many states have developed unique

solid waste landfill programs that reflect specific statewide factors and concerns.

Any evaluation of landfilling options for the management

of WTP residuals should include a careful examination

of state landfill requirements. In many states, the flexibility exists to design and operate landfills according to

site-specific conditions and limitations.

7.1 Landfill Siting

Landfill siting criteria usually address two concerns—

landfill performance and protection of public health and

the environment. Restrictions on siting a landfill in or

near airports, floodplains, wetlands, fault areas, seismic

impact zones, and unstable areas are typically established to protect the landfill’s integrity, thus protecting the

public and minimizing environmental damage should the

landfill fail.

Table 7-1. Solid Waste Landfill Criteria: Monofill for WTP

Residuals and Co-disposal of Residuals With

Nonhousehold Solid Waste (40 CFR Part 257)

Section Subpart

Floodplains 3-1

Endangered species 3-2

Surface water 3-3

Ground water 3-4

Disease 3-6

Air 3-7

Safety 3-8

Table 7-2. Solid Waste Landfill Criteria: Co-disposal of WTP

Residuals With Municipal Solid Wastes (40 CFR

Part 258)

Section Subpart

Location restrictions B

Operating criteria C

Design criteria D

Ground-water monitoring and corrective

action

E

Closure and postclosure care F

Financial assurance criteria G

119
7.1.1 Airports

Scavenger birds such as crows, gulls, and starlings are

the most common aircraft safety concern at landfills,

especially those that receive household solid waste.

Birds are attracted to landfills to nest, roost, and search

for food and water. Therefore, regulations restrict the

construction and operation of landfills in areas surrounding airports. For example, MSWLF criteria (40 CFR Part

258) require that owners/operators of landfills within

10,000 feet of the end of a runway used by turbojet

aircraft and within 5,000 feet of the end of a runway used

only by piston-type aircraft demonstrate that design and

operation of the landfill does not pose a bird hazard to

aircraft. In addition, landfills within a 5-mile radius of the

end of a runway must notify the Federal Aviation Administration of their existence.

For monofills, the issue of airport safety may only be a

concern if co-disposal with household solid waste is

used.

Reducing bird problems at MSWLFs is difficult because

of the tenacity of scavenger birds. Operational practices

that can minimize the attraction of birds to the landfill

include:

• Frequent covering of wastes that provide a source of

food.

• Baling, milling, or shredding food-containing wastes.

• Attempting to eliminate food-type wastes from the

landfill using waste management techniques such as

source separation and composting.

7.1.2 Floodplains

Floodplains are areas inundated by water during a 100-

year flood. These areas usually include lowland and

relatively flat areas next to inland and coastal waters,

including flood-prone offshore islands. Landfills located

in floodplains can restrict river flow during flood events,

reduce the temporary water storage capacity of the

floodplain, or fail, resulting in the washout of solid waste

which can pose a threat to human health and the environment. Therefore, siting landfills in floodplains is usually restricted.

In most cases, these restrictions do not prohibit the

location of a landfill in a floodplain, but a technically

sound demonstration of the design and operational aspects of the landfill that address these concerns is

needed before the applicable regulatory approvals can

be secured. At a minimum, the technical demonstration

requires the landfill owner/operator to identify the

boundaries of the floodplain and its relation to the landfill

and future landfill expansions, and to address engineering considerations. The velocity during base flow and

flood flow conditions and the temporary water storage

capacity of the floodplain must be determined. Modeling

of the effects of the landfill on these velocity and storage

volume characteristics is usually required. The U.S.

Army Corps of Engineers has developed several computer models to assist in the evaluation of these flow and

storage parameters.

Cost-effective methods to adequately protect a landfill

from flood damage have been developed. These methods include using rip-rap and geotextiles to prevent

erosion or other damage to the landfill.

7.1.3 Wetlands

With the concern over increasing loss of wetland resources, the location of landfills in or near wetlands is

severely restricted. The concern relates to the effects of

construction and operation on surface water quality and

fish, wildlife, and other aquatic resources and their habitat, especially endangered or threatened species.

Requirements for the protection of wetlands are found

in numerous federal statutes and orders including:

• Clean Water Act (Sections 401, 402, and 404)

• Rivers and Harbors Act of 1989

• Executive Order 11990, Protection of Wetlands

• Nation Environmental Policy Act (NEPA)

• Migratory Bird Conservation Act

• Fish and Wildlife Coordination Act

• Coastal Zone Management Act

• Wild and Scenic Rivers Act

• National Historic Preservation Act

Besides federal requirements, most states and many

local jurisdictions have requirements to safeguard the

integrity of wetlands. The use of wetlands for the construction of a landfill requires a permit from the U.S.

Army Corps of Engineers, which oversees wetlands

under authority delegated by the U.S. Environmental

Protection Agency (EPA). Usually, owners/operators are

required to demonstrate that no other alternatives exist

to siting a landfill in a wetland, and only under very

limited conditions will landfill construction in wetlands be

approved. Owners/operators are required to develop

extensive plans for mitigating and offsetting wetland

impacts, including the creation of wetlands in other areas.

7.1.4 Fault Areas

Many landfill siting restrictions include a prohibition

against locating a landfill within a certain distance from

an earthquake fault. Seismologists generally believe

that the structural integrity of an engineered unit cannot

be unconditionally guaranteed when built within 200 feet

of a fault along which movement is highly likely to occur.

120
MSWLF criteria (40 CFR Part 258) have established

200 feet as the minimum allowable distance between a

fault and a landfill. Fault areas affected by the 200-foot

limitation are those that have experienced movement in

Holocene time, which is approximately during the last

10,000 years.

The U.S. Geological Survey (USGS) has developed

maps that identify the location of Holocene faults in the

United States. Landfill owners/operators should use

these maps to determine the location of fault areas that

may affect the location of a proposed landfill. A state may

allow the construction of a landfill within a fault area if a

detailed engineering demonstration of the survivability

of the landfill has been developed. Any variance from

these criteria requires a demonstration that no other

practicable waste management alternative exists.

7.1.5 Seismic Impact Zones

In addition to fault areas, most landfill siting restrictions

include consideration of seismic impact zones. A seismic impact zone (40 CFR 258.14) is an area in which a

greater than 10 percent chance exists that the maximum

horizontal acceleration in the bedrock, expressed as a

percentage of the earth’s gravitational pull (g), will exceed 0.10 g in 250 years. Owners/operators can review

the seismic 250-year interval maps in the USGS report

entitled, Probabilistic Estimates of Maximum Acceleration of Velocity in Rock in the Contiguous United States

(USGS, 1982). In addition, the National Earthquake Information Center at the Colorado School of Mines in

Golden, Colorado, can provide seismicity maps of all 50

states.

For landfills to be located in seismic impact zones, an

evaluation of the seismic effects should consider both

foundation soil stability and waste stability under seismic

loading. The evaluation of waste stability for monofills

should be more reliable because of the homogeneous

nature of the waste compared with the heterogeneous

waste properties of municipal solid waste.

While no standard procedures exist for designing landfill

components to withstand seismic events, engineering

evaluations should be conducted on the influence of

local soil conditions on ground response and shaking

intensity, soil settlement, soil liquefaction, and slope

stability during earthquakes.

7.1.6 Unstable Areas

An unstable area is an area susceptible to natural or

manmade events or forces capable of reducing the integrity and performance of all or some of the landfill

components. Factors that contribute to inadequate support are poor foundation conditions; down-slope movement of soil, rock, or debris caused by the influence of

gravity; and sinkholes resulting from Karst terraces underlain by soluble bedrock that may contain extensive

subterranean drainage systems and relatively large subsurface voids.

The construction and operation of landfills in unstable

areas should be avoided. If built, the landfill should be

engineered and constructed to accommodate the conditions that classify the area as unstable. A detailed

geotechnical and geological evaluation of a landfill site

should be conducted to assess the subsurface under

natural and manmade conditions. This evaluation

should analyze the potential for inadequate support for

the structural components of the landfill.

7.2 Landfill Design

As previously discussed, monofills are subject to the

performance-based criteria specified in 40 CFR Part

257. These criteria require that the owner of a monofill,

usually the drinking water utility, demonstrate that the

construction of such a landfill will not adversely affect

surrounding floodplains, endangered species, surface

water, ground water, or air quality, and will not create

disease or safety threats. The criteria do not specify

design considerations that must be followed. Table 7-3

provides some general guidance on the type of information that is necessary before design begins.

The major sludge monofilling methods are trench filling

and area filling. Trench filling can be further subdivided

into narrow trench and wide trench monofilling techniques. Area filling methods include area fill mound, area

fill layer, and diked containment. Sludge solids content,

sludge stability, site hydrogeology (location of ground

water and bedrock), ground slope, and land availability

determine the monofilling method that is selected. The

following sections describe specific aspects of the various sludge monofilling methods. Figure 7-1 highlights

the design consideration for monofills.

A landfill that accepts municipal solid wastes is subject

to the criteria described in 40 CFR Part 258. Under

federal regulations, WTP residuals are subject to the

provisions of Part 258 only if they are co-disposed of in

a landfill that also accepts municipal solid waste. In that

case, the landfill is designated as an MSWLF.

Landfill design under Part 258 can be based on performance or minimum technology standards. These two approaches are described in detail below.

7.2.1 Performance-Based Design Under 

40 CFR Part 258

Performance-based landfill designs are developed so

that the landfill system will result in compliance with

established performance standards. By allowing a designer to use a performance-based approach, site121
specific factors can be considered in the design such as:

• Hydrogeologic characteristics of the landfill and the

surrounding land.

• Volume and physical and chemical characteristics of

the leachate.

• Quality, quantity, and direction of ground-water flow.

• Proximity and withdrawal rate of ground-water users.

• Existing quality of the ground water.

The MSWLF criteria require compliance with performance standards that set maximum contaminant levels

(MCLs) (Table 7-4) not to be exceeded in the ground

water at the relevant point of compliance. The MSWLF

criteria allow for a relevant point of compliance to be

established as far as 150 meters from the waste management unit boundary. In some cases, the relevant

point of compliance must be at the waste management

unit boundary.

Successful application of the performance-based approach, leading to a reliable landfill system design, depends on the designer’s ability to accurately model the

rate of pollutant movement through the landfill system

and the site stratigraphy. Contaminant transport at the

landfill site must be studied carefully to determine the

direction, speed, and concentration of contaminant flow.

Because contaminant transport in ground water can be

very complicated, accurate prediction of contaminant

movement significantly increases the cost of design and

site characterization for an MSWLF (U.S. EPA, 1994a).

7.2.2 Minimum Technology-Based Design

Under 40 CFR Part 258

The MSWLF criteria have established minimum technology-based design standards for landfills that cannot

apply performance-based design standards. Technology-based standards require a composite liner system

consisting of an upper geomembrane liner and a lower

compacted-soil liner. The geomembrane liner minimizes

the exposure of the compacted soil liner to leachate,

thus significantly reducing the volume of leachate reaching the soil liner. Reducing membrane penetration is vital

to controlling the escape of leachate into ground water.

The geomembrane must be at least 30 millimeters thick;

high density polyethylene (HDPE) geomembranes must

be at least 60 millimeters thick. The compacted soil liner

must be at least 2 feet thick and have a hydraulic

conductivity of less than 1 x 10-7 centimeters per second.

Many state MSWLF programs have established alternative technology-based design requirements. Whenever

technology-based designs are being considered, the

state MSWLF program should be contacted to obtain

applicable design requirements.

Table 7-3. Field Investigations for New Information, Landfill

Design (U.S. EPA, 1978b)

General

Information

Specific

Information Method and Equipment

Base Map Property

boundaries

Field survey

Topography and

slopes

Field survey

Surface water Field survey

Utilities Field survey

Roads Field survey

Structures Field survey

Land use Field survey

Vegetation Field survey

Soils Depth Soil boring and compilation of

boring log

Texture Soil sampling and testing via

sedimentation methods (e.g.,

sieves)

Structure Soil sampling and inspection

Bulk density Soil sampling and testing via

gravimetric, gamma ray detection

Porosity Calculation using volume of voids

and total volume

Permeability Soil sampling and testing via

piezometers and lysimeters

Moisture Soil sampling and testing via oven

drying

Ease of

excavation

Test excavation with heavy

equipment

Stability Test excavation of trench and

loading of sidewall of Hueem

stabilimeter

pH Soil sampling and testing via pH

meter

Cation

exchange

capacity

Soil sampling and testing

Bedrock Depth Boring and compilation of boring

log

Type Sampling and inspection

Fractures Field survey

Surface outcrops Field survey

Ground

water

Depth Well installation and initial readings

Seasonal

fluctuations

Well installation and year-round

comparison of readings

Hydraulic

gradient

Multiple well installation and

comparison of readings

Rate of flow Calculation based on permeability

and hydraulic gradient

Quality Ground-water sampling and testing

Uses Field survey via inspection

Climatology Precipitation Rain gauge

Evaporation Class A Evaporation Pan

Temperature Standard thermometer

No. of freezing

days

Minimum-maximum temperature

thermometer

Wind direction Wind arrow

122
Many factors must be considered for a successful

geomembrane design and installation:

• Selection of proper geomembrane materials that consider chemical resistance and biaxial stress-strain

properties.

• Proper subgrade preparation to ensure that the compacted soil liner is smooth and strong enough to

provide continuous support for the geomembrane.

• Proper geomembrane transportation, storage, and

placement.

• Favorable installation conditions, including fair weather,

low winds, and proper temperature.

• Proper geomembrane seaming and seam testing.

• Development and implementation of a reliable and

reasonable construction quality assurance (CQA)

program.

Because of its low hydraulic conductivity characteristics,

clay is usually the soil of choice for compacted soil

liners. Unfortunately, clay is a difficult engineering

material to work with because of its highly moisturedependent physical properties. As a basic landfill liner,

Figure 7-1. Considerations for sludge monofill design (Cornwell et al., 1992).

123
clay must meet certain criteria to protect ground water

from leachate contamination. To meet the requirements

of the MSWLF criteria and to produce a reliable barrier

against leachate movement, the following steps should

be taken during construction of a compacted clay liner:

destroy soil clods, eliminate lift interfaces, conduct

proper compaction, meet moisture-density criteria, and

avoid desiccation.

To minimize holes in a geomembrane liner caused by

product defects, transportation, installation, and seaming, and to produce a reliable soil liner that prevents the

movement of leachate into the ground water, a CQA

program should be developed and implemented (U.S.

EPA, 1993a). The CQA program is a planned system of

activities implemented by the landfill owners or their

representatives, to ensure that the components of the

landfill system are constructed as specified in the design. The CQA program should be developed along with

the landfill design. All involved state and federal regulators should review the CQA program before approvals

and permits for construction are issued.

CQA is different and distinct from construction quality

control (CQC). CQC is a program that the construction

contractor develops and implements to ensure the quality of its work. In contrast, the CQA program is an

additional check of the construction contractor’s work

conducted by the owners of the landfill. Several key

components to a successful CQA program are:

• Responsibility and authority: Landfill owners give

CQA personnel the responsibility and authority to represent their interest and to ensure that the landfill

components meet design specifications.

• Personnel qualifications: The CQA inspector must

have extensive experience and knowledge about the

work performed in the field. A program administered

by the National Institute of Certifying Engineering

Technicians (NICET) gives formal examinations and

certifies CQA inspectors.

• Inspection activities: The CQA program must clearly

define the testing program and acceptance criteria for

significant components of the landfill system. For the

liner system, the CQA program should specify the frequency of testing to be done on the compacted soil and

geomembrane liner, outline the sampling strategy, and

define the specific tests to be performed.

• Sampling strategies: CQA testing is performed using

a combination of statistical and judgmental sampling

strategies. Typical statistical sampling strategies include defined interval testing, such as one destructive

seam test per 5,000 feet of geomembrane seam or

one moisture/density test per 5,000 cubic yards of

soil liner. Judgmental testing allows the CQA inspector to call for testing when the quality of workmanship

is suspect.

• Documentation: Before a permit to operate the landfill

is issued, most states now require documentation

that a CQA program was performed. All CQA activities must be clearly documented so that a third party

can understand and verify the testing and inspection

program.

Another requirement of technology-based systems in

the design and construction of a leachate collection

system is to collect leachate and convey it out of the

landfill. The MSWLF criteria require this system to ensure that less than 30 centimeters of leachate accumulates over the composite liner system to minimize

hydraulic head and possible contamination of the

ground water. When a leachate collection system is

designed and constructed, the following components

should be included:

• Area collector: the drain that covers the liner and

collects the leachate.

Table 7-4. Maximum Contaminant Levels in Uppermost

Aquifer at Relevant Point of Compliance (Fed.

Reg., October 9, 1991)

Parameter MCL (mg/L)

Arsenic 0.05

Barium 1.0

Benzene 0.005

Cadmium 0.01

Carbon tetrachloride 0.005

Chromium (hexavalent) 0.05

2,4-Dichlorophenoxy acetic acid 0.1

1,4-Dichlorobenzene 0.075

1,2-Dichloroethane 0.005

1,1-Dichloroethylene 0.007

Endrin 0.0002

Fluoride 4.0

Lindane 0.004

Lead 0.05

Mercury 0.002

Methoxychlor 0.1

Nitrate 10.0

Selenium 0.01

Silver 0.05

Toxaphene 0.005

1,1,1-Trichloromethane 0.2

Trichloroethylene 0.005

2,4,5-Trichlorophenoxy acetic acid 0.01

Vinyl chloride 0.002

124
• Collection laterals: the pipe network that drains the

area collector.

• Sump: the low point where leachate exits the landfill.

• Stormwater/Leachate separation system: a system

for minimizing leachate generation and possible treatment costs.

These components must be designed to handle larger

leachate flows associated with initial operations and to

resist problems such as biological clogging that can

destroy the long-term performance and flow capacity of

the system.

7.3 Landfill Operations

To meet the regulatory requirements for liners and

leachate collection systems, landfill design and construction have become increasingly complex. Thus, the

integrity of landfill components can be threatened by

careless or inappropriate operations. Facility operators

should be fully aware of landfill operational requirements

and the reasons for these requirements to ensure that

the landfill performs as designed, especially in situations

where the reason for some operational procedures may

not be readily apparent. At the same time, landfill designers and regulators should obtain feedback from

landfill operators on day-to-day operational requirements. A complex, sophisticated design that cannot

be operated in the field will not achieve its intended

purpose.

Operational requirements for landfills are designed to

ensure the safety of people on the landfill, including

landfill operators, waste haulers, and the public, and to

protect the environment. MSWLF criteria require the

following list of measures to be implemented at all

landfills:

• Exclude hazardous waste and polychlorinated

biphenyls (PCBs).

• Provide daily cover.

• Control onsite disease vectors.

• Provide routine methane monitoring.

• Eliminate most open burning.

• Control public access.

• Institute run-on and runoff controls.

• Control discharges to surface waters.

• Eliminate the disposal of most liquid wastes.

• Keep records that document implementation of 

operational requirements.

7.4 Metal Content Considerations

The most prominent environmental consideration in the

siting, construction, and operation of a residuals monofill

under 40 CFR Part 257 is the potential for ground-water

contamination through the leaching of metal constituents in the residuals. Assessing the potential for contamination involves determining the metals concentration

in the residuals and the degree to which these metals

can be expected to mobilize into ground water.

The major sources of metals are raw water and treatment chemicals. Table 6-2 presents the findings of one

investigation into the metals content of water treatment

chemicals (Dixon et al., 1988). Aluminum is, of course,

a major constituent of alum coagulant, and iron is a

major constituent in ferric chloride coagulant; these metals predominate in WTP residuals as a result. The data

in Table 6-2 indicate that many heavy metal impurities

can exist in coagulants and in other treatment chemicals. Raw water sources also can contain metals naturally or because of source contamination. Raw water

metals largely wind up in WTP residuals.

Metals levels in both treatment chemicals or raw water

can vary, especially in the case of raw water. Because

metals concentrations in WTP residuals vary widely at

different water treatment facilities, a general characterization of metals concentrations in WTP residuals is

difficult to make. Such a characterization is complicated

by the lack of comprehensive surveys of metals concentration in WTP residuals.

Nonetheless, a rough characterization of typical WTP

residuals metals concentrations is presented in Table

7-5. The number of samples from which each range was

compiled should be noted when interpreting the ranges.

The American Water Works Service Co. survey of 19

water plants in Pennsylvania covered only seven metals—cadmium, chromium, copper, lead, mercury, nickel,

and zinc (Dixon et al., 1988)—which accounts for the

disparity in sample set sizes.

Typical metals concentration ranges in natural soils and

sewage sludge are provided in Table 7-5 for comparison

with WTP residuals (Dragun, 1988). Median concentrations for sewage sludge are also presented (U.S. EPA,

1984b). The concentration data for both natural soils

and sewage sludge are based on more comprehensive

information than could be compiled for WTP residuals.

The three substances covered in Table 7-5 are roughly

comparable. WTP residuals usually contain much lower

metals levels than typical sewage sludge contains and

often levels are very similar to those found in natural

soils.

125
7.4.1 Classification as Hazardous or

Nonhazardous Waste

The total metals concentrations in a material do not, by

themselves, give a good indication of the potential for a

material to contaminate ground water with metals in a

landfill situation; the degree to which the metal constituents readily solubilize into ground water is an equally

important factor. Laboratory extraction testing is one

method to assess the extent to which a material’s metals

will readily solubilize.

As part of the RCRA Subtitle C regulations, EPA has

defined four extrinsic characteristics, any one of which

can qualify a waste as hazardous. The four characteristics are reactivity, corrosivity, ignitability, and toxicity.

Since the inception of the RCRA regulations, toxicity has

been defined in terms of extraction testing results. Prior

to May 1990, the extraction procedure (EP) toxicity test

was the standard test for determining toxicity. Since May

1990, the standard test has been the toxicity characteristics leaching procedure (TCLP) test. Both tests involve the preparation of an aqueous extract from a

waste sample, which is in turn analyzed for a number of

toxic parameters including metals (the same eight in

each test) and organic compounds.

Although extraction testing is the means by which a

substance is determined to be toxic, according to the

RCRA regulations, it is also a means for determining the

mobility of toxic analytes in a waste that is not toxic. For

WTP residuals, failure of the TCLP test is not likely

because the metals content of most residuals samples

is low enough that even if 100 percent of each metal

were extracted into the extraction fluid, the TCLP concentration limits would not be exceeded (Cornwell et al.,

1992).

Table 7-6 demonstrates this point by presenting the

TCLP concentrations that would result from 100 percent

extraction of the metals in a residuals sample having

metals concentrations equal to the maximums given in

Table 7-5. These TCLP concentrations are based on the

assumption of a 35 percent solids sample. As part of the

TCLP test, samples are subjected to pressure filtration

to remove excess water. Dewatered coagulant residuals

will not yield any water when subjected to pressure

filtration and will, therefore, be tested at their dewatered

solids concentration. Even 100 percent extraction of the

TCLP metals from a hypothetical dewatered residuals

sample that contained all the maximum metals concentrations given in Table 7-5 would not cause an exceedance of any of the TCLP regulatory limits. This

clearly indicates that WTP residuals should rarely, if

ever, qualify as toxic under RCRA regulations.

This conclusion is supported by actual extraction test

results for WTP residuals. Table 7-7 presents actual

Table 7-5. Comparison of Metals Concentrations in WTP Residuals, Natural Soils, and Sewage Sludge

Metals

Number of WTP

Residuals

Samples Included

in Compilation

Range of

Concentrations in

WTP Residuals

(mg/kg, dry)a

Average

Concentration in

Compiled WTP

Residuals

Samples

(mg/kg, dry)a

Typical Range of

Concentrations

in Natural Soils

(mg/kg, dry)b

Typical Range of

Concentrations in

Sewage Sludge

(mg/kg, dry)c

Median

Concentration in

Sewage Sludge

(mg/kg, dry)c

Aluminum 7 10,000–170,000 72,729 10,000–300,000 N/A N/A

Arsenic 6 5.7–36 20 1.0–40 1.1–230 10

Barium 6 30–333 131 100–2,500 N/A N/A

Cadmium 24 0–16 3.3 0.01–7.0 1–3,410 10

Chromium 26 6.7–200 59 5.0–3,000 10–99,000 500

Copper 25 7–1,300 176 2.0–100 84–17,000 800

Iron 4 15,200–79,500 51,400 7,000–550,000 1,000–154,000 17,000

Lead 26 1–100 50 2.0–200 13–26,000 500

Manganese 5 68–4,800 1,453 100–4,000 32–9,870 260

Mercury 24 0–9.8 0.9 0.01–0.08 0.6–56 6

Nickel 23 12–1,319 209 5.0–1,000 2–5,300 80

Selenium 5 0–36 8 0.1–2.0 1.7–17.2 5

Silver 4 0–2 1 0.1–5.0 N/A N/A

Zinc 25 6.2–3,325 598 10–300 101–49,000 1,700

a Cornwell et al., 1992; Dixon et al., 1988.

b Dragun, 1988.

c

 U.S. EPA, 1984b.

Key

N/A = not analyzed.

126
results for EP toxicity tests performed on several WTP

residuals samples and Table 7-8 presents TCLP results

for WTP residuals from five facilities. In all cases, no

TCLP regulatory limits were even approached (Cornwell

et al., 1992). In fact, TCLP metals often cannot be

detected in extraction fluid because concentrations are

so low.

Because the detection limits of metals are often encountered in such testing, it is difficult to measure the percentage of a metal’s original amount. The results from

pilot lysimeter testing, however, often can be expressed

in terms of percentages of metals extracted.

7.4.2 Mobility of Trace Metals

Lysimeter testing can provide additional insight into the

mobility of metals from WTP residuals. A recent study

used the lysimeter setup shown in Figures 7-2A to 7-2C

to test the leachability of metals from three residuals

samples (Cornwell et al., 1992). The residuals used in

this lysimeter test came from three WTPs. Two alum

residuals samples were used, one from the Williams

WTP in Durham, North Carolina, and one from the

Chesapeake WTP in Chesapeake, Virginia. The third

residuals sample was a ferric chloride sample from the

Aldrich WTP operated by the Pennsylvania-American

Water Co. TCLP test results are presented in Table 7-8.

The study lasted 24 weeks. Each week, 44.1 liters of

simulated rainwater were pumped over a 24-hour period

into each lysimeter column. The simulated rainwater

was made in the laboratory, having a pH of 4.5 and

containing the constituents listed in Table 7-9. A perforated plate was used to hold the rainwater and allow it

Table 7-6. Worst Case TCLP Results Using Maximum Metals

Concentrations in Table 7-5, Compared With TCLP

Regulatory Limits (Cornwell et al., 1992)

Metal

Maximum Possible

Concentration in

Extraction Fluid

(mg/L)

TCLP

Regulatory

Limit (mg/L)

Arsenic 0.63 5.0

Barium 5.83 100.0

Cadmium 0.28 1.0

Chromium 3.50 5.0

Lead 1.75 5.0

Mercury 0.17 0.2

Selenium 0.63 1.0

Silver 0.04 5.0

Table 7-7. EP Toxicity Test Results for Alum Residuals (Cornwell et al., 1992)

Contaminant

Regulatory

Threshold

(mg/L)

Saltonstall, CT

(mg/L)

West River,

CT (mg/L)

City of

Chesapeake,

VA (mg/L)

American Water Works

Service Co. 66-Plant

Survey (mg/L)

Arsenic 5.0 <0.01 <0.01 <0.003 <0.2–0.4

Barium 100.0 0.21 0.1 <0.1 <0.1–34.0

Cadmium 1.0 <0.005 <0.005 0.005 <0.005–0.06

Chromium 5.0 <0.01 <0.01 <0.05 <0.1–3.8

Lead 5.0 <0.01 <0.001 <0.001 <0.0004–0.003

Mercury 0.2 <0.001 <0.001 <0.001 <0.0004–0.003

Selenium 1.0 0.09 <0.01 <0.007 <0.001—0.08

Table 7-8. TCLP Results for Five Water Treatment Plant Coagulant Residuals (Cornwell et al., 1992)

Contaminant

Regulatory

Threshold

Level (mg/L)

Contra Costa, CA,

Water District

Alum Sludge (mg/L)

Phoenix, AZ,

Alum Sludge

(mg/L)

Durham, NC,

Alum Sludge

(mg/L)

Chesapeake,

VA, Alum

Sludge (mg/L)

Aldrich, PA,

WTP Ferric

Sludge (mg/L)

Arsenic 5.0 0.04 <0.3 0.0088 0.0023 0.006

Barium 100.0 1.1 1.1 1 <1 <2

Cadmium 1.0 <0.05 <0.02 <0.02 <0.02 0.02

Chromium 5.0 0.06 <0.04 <0.1 <0.1 <0.1

Lead 5.0 <0.2 <0.5 <0.1 <0.1 <0.1

Mercury 0.2 <0.001 <0.01 <0.002 <0.002 <0.002

Selenium 1.0 <0.02 <1 <1 <0.002 <0.002

Silver 5.0 <0.05 <0.01 <0.1 <0.1 <0.1

127
to drip to the surface of each lysimeter in a manner

similar to rainfall. Residuals were placed loosely in each

lysimeter column to allow the rainwater to percolate

through. Because residuals are a nearly impermeable

material, had they been compacted the water would

have formed a pond at the top of the columns and not

drained through. Rainwater was allowed to sit in the

columns for a week before being drained for sampling.

The total amount of rainwater applied to each lysimeter

was equivalent to 450 inches of rainfall, about 12 years

of rain in Virginia. The total metals concentrations for

these residuals are presented in Table 7-10.

Table 7-11 summarizes the results from the lysimeter

study in terms of the percentages of each metal leached.

Manganese was the only metal that leached at a rate of

over 1 percent in all three residuals samples. This

lysimeter study clearly indicates that not only do WTP

residuals generally contain small concentrations of metals, but only small portions of these metals tend to leach.

In the test, aluminum, barium, chromium, lead, and silver did not leach measurably from any of the residuals

samples. Cadmium and nickel leached only from the

ferric chloride residuals. Arsenic, copper, iron, manganese, and zinc leached from all the residuals samples.

Table 7-12 presents the maximum leachate concentrations of each metal measured as part of the lysimeter

study. As indicated in the table, no primary MCLs were

exceeded in any of the leachate samples, although the

secondary MCLs (SMCLs) for iron and manganese

were exceeded. The iron SMCL was exceeded in the

leachate of one of the alum residuals samples and the

ferric residuals sample, while the manganese SMCL

Figure 7-2A. Compaction curve, Ferric Sludge 3 (Cornwell et

al., 1992).

Figure 7-2B. Compaction curve, Alum Sludge 1 (Cornwell et al.,

1992).

Table 7-9. Chemical Constituents of Synthetic Rainwater

(Cornwell et al., 1992)

Constituent Concentration (mg/L)

NH3 0.045

CA2+ 0.14 

Mg2+ 0.073

Na+ 0.46 

K+ 0.078

Cl- 1.63 

N3-5 0.036

SO42- 1.54 

pH 4.5 

Figure 7-2C. Compaction curve, Alum Sludge 2 (Cornwell et al.,

1992).

128
was exceeded in the leachate from all three residuals

samples. These SMCLs are not a part of the list of

indicator parameters for ground-water protection monitoring given in federal municipal landfill regulations.

Leachate pH was monitored in addition to metals concentrations as part of the lysimeter study. pH has a

crucial impact on the leaching of metals from wastes.

The simulated rainwater applied to each lysimeter had

a pH of 4.5; the pH of leachate samples was considerably higher than this. The leachate from Durham’s alum

residuals had a pH range of 5.5 to 6. The other two

residuals samples produced leachate in the pH range of

6.5 to 7. pH buffering was exhibited by all three residuals

samples. This quality was probably a key factor in keeping the solubility of metals low and limiting metal

leaching.

The testing of metals concentrations and mobility in

WTP residuals indicates that leachate from WTP residuals should not, in most cases, cause ground water to

exceed drinking water MCLs for metals. The permeability of the residuals was not accounted for, however, in

any of the extraction or lysimeter tests. As part of an

American Water Works Association Research Foundation (AWWARF) landfill study, the Unified Soil Classification System was used to determine that the residuals

samples tested exhibited the characteristics of a CH soil

group (i.e., inorganic clay of high plasticity). The permeability of this class of soil is typically less than 2 x 10-8

inches per second. This is a lower permeability than that

of typical clay liners, which must be at least 4 x 10-8

inches per second.

7.5 Dewatering

Before landfilling, WTP residuals are often dewatered to

reduce the liquid content and increase the solids concentration. This is necessary so that the residuals will

have a high enough solids content to qualify for disposal

in a landfill. Sludge dewatering methods are both mechanical and nonmechanical. Mechanical dewatering

methods use vacuum or pressure to remove the liquid

Table 7-12. Maximum Metals Concentrations in Lysimeter

Leachate Compared With MCLs (Cornwell et al.,

1992)

Metal

Durham,

NC, Alum

Sludge 1

(mg/L)

Chesapeake,

VA, Alum

Sludge 2

(mg/L)

Aldrich,

PA, WTP

Ferric

Sludge 3

(mg/L)

MCL

(mg/L)

Arsenic 0.0035 0.0019 0.0366 0.05 

Cadmium — — 0.01 0.01 

Copper 0.12 0.03 — 1.0 

Iron 2.00 0.30 1.00 0.3a

 

Manganese 9.50 0.36 22.80 0.05a

Nickel — — 0.06 N/A 

Zinc 0.036 0.017 0.373 5.0 

a Secondary MCL.

Table 7-10. Total Metals Analysis for Sludges Used in

Leaching Research (Cornwell et al., 1992)

Metal

Durham,

NC, Alum

Sludge 1

(mg/kg

dry weight)

Chesapeake, 

VA, Alum

Sludge 2

(mg/kg dry

weight)

Aldrich, PA,

WTP Ferric

Sludge 3

(mg/kg dry

weight) 

Aluminum 107,000 123,000 28,600

Arsenic 25.0 32.0 9.2

Barium 30 <30 230

Cadmium 1 1 2

Chromium 120 130 50

Copper 168 16 52

Iron 48,500 15,200 79,500

Lead 11 9 40

Manganese 1,180 233 4,800

Mercury 0.1 <0.1 0.2

Nickel 24 23 131

Selenium <2 <2 <2

Silver <2 <2 <2

Zinc 91.7 393 781

Table 7-11. Leaching of Metals in Lysimeter Test (Cornwell et

al., 1992)

Percentage of Total Metal Present That

Actually Leached

Metal

Alum

Sludge 1

Alum

Sludge 2

Ferric

Sludge 3

Aluminuma 0.00 0.00 0.00

Arsenic 0.05 0.05 2.77

Barium 0.00 0.00 0.00

Cadmium 0.00 0.00 2.03

Chromium 0.00 0.00 0.00

Coppera 0.12 0.42 0.08

Irona 0.03 0.05 0.01

Lead 0.00 0.00 0.00

Manganesea 12.48 3.38 2.45

Nickela 0.00 0.00 0.03

Seleniumb — 0.00 —

Silver 0.00 0.00 0.00

Zinca 0.13 0.05 0.08

a These metals were not analyzed in a standard TCLP test.

b Selenium leaching could not be quantified.

129
portion of the sludge, whereas nonmechanical methods

depend on gravity and evaporation to reduce the undesired liquid in the sludge.

7.5.1 Mechanical Dewatering

The equipment used most frequently for mechanical

dewatering includes the vacuum filter, continuous belt

filter press, plate pressure filter, and centrifuge. An essential element of the vacuum filter is the rotating cylindrical drum. Lying on its side, the drum rotates around

its lengthwise axis with its lower part submerged in a

tank of water sludge. Its periphery surface is covered

with a permeable fabric or other filter material. A vacuum

is applied to suck water through the fabric. Solids left on

the drum surface, called filter cake, are then removed

using a scraper blade.

The continuous belt filter press is equipped with two filter

belts, one on top of the other, which wind through a

series of rolls. The water sludge contained between the

two belts is subjected to bending and shearing action as

the belts continuously move around the rolls. This bending and shearing action causes the sludge to drain and

form a sludge cake.

A plate pressure filter contains a series of recessed

plates covered with a filter cloth. The water treatment

sludges are pumped between the plates, and the liquid

seeps through the filter cloth, leaving behind the sludge

cake. When the spaces between the plates are filled,

the sludge cakes are removed.

The centrifuge method of dewatering uses centrifugal

force to generate pore water pressure in the sludge.

When pore water pressure develops, water drains out

of the sludge, leaving the sludge cake in a machine. The

centrifuge rotating speed typically ranges between 800

and 2,000 rotations per minute. With other parameters

being equal, the faster the rotating speed, the greater

dewatering power.

7.5.2 Nonmechanical Methods

In nonmechanical dewatering, WTP residuals are

spread out in a lagoon or a sand drying bed, exposed to

the air, and allowed to dry by evaporation.

Lagoons are large basins that are either enclosed with

earthen embankments or excavated for WTP residuals

deposits. They can be used for either dewatering or

residuals storage. Dewatering lagoons have pervious

bottoms such as sand layers with underdrains to facilitate drainage; storage lagoons must have impermeable

bottoms to protect the ground water. Storage lagoons

generally are equipped with a decanting facility such as

pumps to remove the standing water.

Sand drying beds are shallow basins with sand bottoms

and underdrain systems. They are also equipped with

facilities for decanting the standing water. Sand drying

beds perform the same function as dewatering lagoons;

the main difference is the thickness of deposited sludge

and the frequency of deposition. Sand drying beds receive thinner layers and more frequent deposition of

sludges than dewatering lagoons receive.

The degree of dewatering depends on the effectiveness

of the dewatering method and the water-holding capacity of the sludge. Because the water-holding capacity of

sludges can vary, the obtainable solids concentration of

sludge cakes also varies by both dewatering methods

and sludge types.

7.6 Physical Characteristics of Water

Sludges

The physical characteristics of water sludges that need

to be addressed in any effective landfill plan include

plasticity, compaction behavior, compressibility, and

shear strength. Of these, plasticity characteristics affect

sludge handleability. Compaction data, which provide

moisture-density curves, are needed for landfill construction control. Compressibility data are necessary to

determine landfill settlement. Shear strength affects the

landfill stability, which controls the maximum height and

slope of a stable landfill and the ability of the landfill to

support heavy equipment.

7.6.1 Plasticity

The plasticity characteristics of water sludges can be

evaluated from the value of Atterberg limits, especially

liquid limit and plasticity index. The plasticity index

equals the difference between the liquid limit and the

plastic limit. Both liquid and plastic limits are expressed

as water content in percentage. The water content is

defined as the ratio between the weights of water and

solid phases, which can be related to the solids content

as shown in Equation 7-1, in which the solids content is

the ratio between the weights of solids and the sludge

mass expressed in percentage:

This equation can be reduced to:

The Atterberg limits of WTP residuals vary greatly with

the type of sludge. The lime/alum/polyamine coagulant

water sludge of the Jersey City Water Treatment Plant

has only slight to little plasticity (Raghu et al., 1987). As

water

content (%) = 





100

solids content (%)

−1





(100)

(Eq. 7-1)

water

content (%) = 1 −







solids content(%)

100







solids content (%)

100 × 100

(Eq. 7-2)

130
shown in Table 7-13, however, the liquid limit, plastic

limit, and plasticity index of Alum Sludge 1, Alum Sludge

2, and Ferric Sludge 3 are very high (Cornwell et al.,

1992). Alum Sludge 1 was produced from treatment of

medium-color, medium-turbidity raw water, dewatered

using the sand drying bed method and obtained from

Williams Water Treatment Plant in Durham, North Carolina. Alum Sludge 2 was generated from high-color,

low-turbidity raw water, dewatered using the sand drying

bed method and obtained from Chesapeake Water

Treatment Plant in Chesapeake, Virginia. The ferric

sludge was generated from medium-color, medium-turbidity raw water, dewatered using the lagoon method

and obtained from Aldrich Water Treatment Plant in

Pittsburgh, Pennsylvania.

Materials with a high plasticity index are considered

highly plastic and difficult to handle. Thus, of these four

different sludges, the lime/alum/polyamine coagulant

sludge is the easiest to handle, and Alum Sludge 2 is

the most difficult.

7.6.2 Compaction Behavior

Compaction data are used to establish the degree of

sludge compaction necessary to increase stability, decrease permeability, and enhance resistance to erosion.

The shape of the moisture-density relation (compaction

curve) varies greatly with sludge type. For the lime/alum/

polyamine coagulant sludge (Raghu et al., 1987), the

moisture-density curve of the modified Proctor compaction exhibits the typical one-hump shape; the optimal

water content and maximum dry unit weight equal 65

percent and 51 pounds per cubic foot (lb/ft3

), respectively. The moisture-density curve of a coagulant sludge

under the Standard Proctor compaction also exhibits the

typical one-hump shape, with the optimal water content

and maximum dry unit weight of 17 percent and 105

lb/ft3

 (16.5 kN/m3), respectively. Of the three coagulant

sludges shown in Table 7-13, however, only the ferric

sludge exhibits the one-hump shape; the optimal water

content and maximum dry unit weight of the Standard

Proctor equal 45 percent and 72 lb/ft3

 (11.3 kN/m3),

respectively.

Both Alum Sludges 1 and 2 show a decrease in dry unit

weight with increasing water content without a peak

formation. Meanwhile, for Alum Sludge 2, admixing a

bulking agent—a slaked lime, a class C fly ash, or a

natural soil—insignificantly affects the moisture-density

relation. The amount of additive equals 60 percent by

dry weight of the sludge. The moisture-density relations

of these three sludges are plotted in terms of percent

solids concentration versus dry unit weight in Figure 7-3.

Each curve is accompanied by the zero-air-void curve

(ZAVC) that outlines the upper limit of the curve regardless of the compaction effort used. The ZAVC is obtained from Equation 7-3:

dry

unit

weight

= (specific gravity) (unit weight of water)

(1 + specific gravity)







water content (%)

100







(Eq. 7-3)

For landfill construction, the sludge should be compacted to the highest possible density that can be accomplished only when the amount of water in the sludge

is at the optimal water content. For Alum Sludges 1 and

2, however, the compaction water content should be as

low as possible.

7.6.3 Compressibility

In the landfill, under the sustained weight of the overlying mass, sludge may undergo volume reduction from

the extrusion of water. As a result, the landfill may settle,

creating air space for depositing additional sludge on top

of the landfill. To determine the amount of landfill settlement, the compressibility property of the sludge is

required.

Table 7-13. Liquid Limit, Plastic Limit, and Plasticity Index of

Water Sludges (Cornwell et al., 1992)

Water Sludge

Liquid

Limit (%)

Plastic

Limit (%)

Plasticity 

Index

Alum Sludge 1 423 137 286

Alum Sludge 2 550 239 311

Ferric Sludge 3 108 47 61

Consolidation Pressure (kPa)

Void Ratio (e)

Ferric Sludge 3

Alum Sludge 1

Alum Sludge 2

Figure 7-3. Consolidation curves of Alum Sludges 1 and 2 and

Ferric Sludge 3 (Wang et al., 1992b).

131
Normally, the compressibility property is obtained from

the consolidation test. The test results can be analyzed

using the conventional small-strain theory or the finitestrain theory of consolidation. Although the small-strain

theory is more widely used, the resulting analyses are

not as accurate as those that use the finite-strain theory.

Based on the traditional small-strain theory, Raghu et al.

(1987) reported a compression index (Cc) of approximately 0.3 for their lime/alum/polyamine coagulant

sludge. The compression index (Cc) is computed from

the slope of the field compression curve, which relates

the void ratio (e) with the logarithm-of-consolidation

pressure (p), namely, e versus log p curve. The void ratio

is the ratio between the volume of void and the volume

of solid phase of the sludge.

The e versus log p curves of Alum Sludges 1 and 2 and

Ferric Sludge 3 are shown in Figure 7-3 (Wang et al.,

1992b). From these curves, Cc equals 6.69, 5.29, and

1.99 for Alum Sludge 1, Alum Sludge 2, and Ferric

Sludge 3, respectively. Under the same sustained pressure, the higher the Cc value is, the greater the landfill

settlement will be. Thus, of these four sludges, it can be

expected that the landfill of Alum Sludge 1 will settle

most, while the lime/alum/polyamine sludge will settle

least, other factors being equal.

From Figure 7-3, the swelling index (Cs) can also be

obtained. The value of Cs is computed from the slope of

the swelling (or rebound) portion of the graph and is

normally used to determine the amount of landfill rebound caused by reduction of the sustained weight. The

values of Cs are 0.17, 0.15, and 0.03 for Alum Sludge 1,

Alum Sludge 2, and Ferric Sludge 3, respectively. Thus,

of these three sludges, the amount of landfill rebound

will be greatest for Alum Sludge 1 and smallest for Ferric

Sludge 3.

The effect of a bulking agent on the e versus log p

relation of Alum Sludge 1 is illustrated in Figure 7-4

(Wang et al., 1992a). As shown, admixing of a bulking

agent results in a lower e versus log p curve. To consider

the effect of pozzolanic reaction in the lime and fly ash

treated sludge, these two treated sludges were curved

to two different durations: 2 weeks and 4 weeks. The

compression indexes of these treated sludges are 4.04

for the soil-treated, 3.56 for the lime-treated and 2-week

cured, 4.34 for the lime-treated and 4-week cured, 3.70

for the fly ash-treated and 2-week cured, and 3.93 for

the fly ash-treated and 4-week cured.

7.6.4 Shear Strength

Shear strength is essential for the design and analysis

of landfill stability with or without heavy equipment operating on the landfill. Different testing methods are

available for the determinations of shear strength, including cone penetration test, vane shear test, and

triaxial compression test. The first two methods normally

are used to measure undrained shear strength, while the

third method is used to determine both undrained and

drained strengths. The undrained shear strength is for

use when there is insufficient time for the sludge to drain

during shearing. Under the undrained loading condition,

excess pore water pressure develops, reducing the capacity of sludge particles to resist shear and resulting in

a lower shear strength. On the other hand, the drained

shear strength is for when, during shearing, the sludge

drains completely without developing excess pore water

pressure. The sludge particles more effectively resist

shear, resulting in a higher shear strength. The undrained shear strength is used for analysis of landfill

stability during or immediately after construction; the

drained shear strength is needed in the analysis of

long-term stability of the landfill.

Untreated Alum Sludge 1

Treated With 


Lime

With Soil

With Fly Ash

Consolidation Pressure (TSF)

Void Ratio 17.5


15.0


12.5


10.0


7.5


5.0


2.5

10 10 1 10 -2 -1

Figure 7-4. Void ratio versus consolidation pressure of treated

and untreated Alum Sludge 1 (Wang et al., 1992a).

Alum Sludge 1

Alum 


Sludge 2

Ferric


Sludge 3

Solids Content (%)

Undrained Shear Strength (KN/M2

)

30


25


20


15


10


5


0

10 20 30 40 50 60

Figure 7-5. Strength versus solids content for Alum Sludges 1

and 2 (Cornwell et al., 1992).

132
The shear strength property of water treatment sludges

varies not only with the type and nature of the sludge

but with solids content as well. The rate and amount of

strength increase with increasing solids content also

differ for various sludge types. Figure 7-5 shows the

undrained shear strength versus solids content curves

for Alum Sludges 1 and 2 and Ferric Sludge 3. As

shown, the shear strength increases very slowly in the

beginning and should approach a constant value at very

high solids content for each sludge, although the later

portion of the curve is not seen within the range of

conditions investigated. The solids content required for

a given shear strength is highest for Alum Sludge 1 and

lowest for Ferric Sludge 3. Also, at a given solids content, the shear strength is greatest for Alum Sludge 1

and least for Ferric Sludge 3.

The shear strength parameters and pore water pressure

parameter at failure (Af) obtained from the triaxial compression test are summarized in Table 7-14 (Wang et al.,

1992b). Also included is the range of initial solids content

of the test specimens. The initial solids content of each

sludge is near the value required for ease in test specimen preparation. The Af

 value of 0.75 to 0.79 falls within

the range for normally consolidated clays or sensitive

clays. The effective internal friction angle (ø), however,

is unusually high when compared with highly plastic

clay soils.

Water treatment sludges exhibit rheotropic and thixotropic behavior; namely, sludges undergo strength reduction from disturbance or remolding, and after

disturbance the strength increases with curing time without a change in solids content. The reduction in sludge

strength from disturbance has been experienced by

workers who operated a hydraulic excavator on a trench

fill of Chesapeake sludge (Alum Sludge 1). According to

the machine operator, the sludge that initially was stable

enough to operate the machine on became weak and

unstable over time (Cornwell et al., 1992).

The rate and amount of strength increase from curing

depends on sludge type and solids content. The increase of undrained shear strength at different solids

contents for the ferric sludge is shown in Figure 7-6.

From these strength data, the strength-gain ratios are

computed for different sludges at different solids contents and are shown in Table 7-15. The strength-gain

ratio is the ratio between cured strength and remolded

strength. The data in Table 7-15 indicate that the

strength-gain ratio appears first to increase, then to

decrease with decreasing solids content. Thus, there is

an optimal solids content for the strength-gain ratio to

reach the maximum.

The rheotropic and thixotropic behavior of sludge should

be properly considered in the analysis, design, and

construction of sludge landfills. Because construction

operation inevitably causes sludge remolding, it is appropriate to use the remolded undrained shear strength

to design the landfill and also to analyze the stability of

the landfill during construction. After construction, the

shear strength of sludge increases with time from thixoTable 7-14. Shear Strength Parameters of Test Sludges

(Wang et al., 1992b)

Total Stress Effective Stress

Initial

Content

c

(kPa)

ø

(°)

c′

(kPa)

ø′

(°)

Pore

Pressure

Parameter

Af

Alum

Sludge 1

13.4–15.5 4.14 19.3 6.89 42.3 0.75

Alum

Sludge 2

24.3–26.3 4.83 19.0 8.27 44.0 0.77

Ferric

Sludge 3

36.2–40.4 8.27 17.5 8.27 42.8 0.79

Curing Time (Day)

33.5%

35.4%

36.8%

SC = 41.2%

Undrained Shear Strength (KN/M2

)

Figure 7-6. Strength versus curing time at various solids contents for ferric sludge (Wang et al., 1992b).

Table 7-15. Remolded and Cured Undrained Strengths and

Strength Gain Ratio (Wang et al., 1992b)

Solids

Content

(%)

Remolded

Strength

(kPa)

Cured

Strength

(kPa)

Strength

Gain

Ratio

Alum Sludge 1 16.3 2.21 12.96 5.9

13.6 1.10 7.65 7.0

11.9 0.48 3.86 8.0

11.1 0.34 1.93 5.7

Alum Sludge 2 19.3 2.76 17.93 6.5

18.6 2.07 115.17 7.3

Ferric Sludge 3 41.2 2.34 7.24 3.1

36.8 1.03 3.45 3.4

35.4 0.83 1.86 2.2

33.5 0.62 1.24 2.0

133
tropic hardening and consolidation. Consequently, the

landfill becomes more stable.

Admixing a bulking agent enhances the shear strength

of the sludge. Figure 7-7 shows that for Alum Sludge 2

without curing, the degree of strength improvement varies with the type of bulking agent, being greatest for the

slaked lime, fly ash, and finally the soil. Furthermore, for

each bulking agent, strength increases with increasing

treatment level (Cornwell et al., 1992). The strength

increase from curing also differs for various bulking

agents, as illustrated in Figure 7-8 (Cornwell et al., 1992).

Although the ultimate cured strength is greatest for lime,

then fly ash, soil, and the untreated sludge, the strengthgain ratio is in reverse order, as shown in Table 7-16.

The data in Table 7-16 indicate a drastic increase in the

remolded shear strength of Alum Sludge due to treatment, by a factor of more than four times for lime, almost

three times for fly ash, and about one-and-a-half times

for soil. Such an improvement in shear strength makes

it possible to construct a higher landfill with a steeper

slope. Additionally, admixing stabilization makes the

sludges less plastic and therefore easier to handle.

Using the shear strength data in landfill analysis and

design has been demonstrated by Cornwell et al. (1992)

and Wang et al. (1992b). Specifically, the strength data

were used to make a hypothetical monofill design for the

three different sludges with varying solids concentration.

The maximum fill height for a desired slope angle or the

maximum slope angle for a desired fill height were

determined. The hypothetical monofill has a uniform

slope with a constant slope angle from the toe to the top

of the slope. The top of the landfill is level, extends very

far, and carries no surface loading. Meanwhile, the landfill is supported on very large, firm, level ground, which

allows it to be treated as a two-dimensional problem.

Figure 7-9 shows landfill height versus the ratio between

shear strength and wet unit weight of sludge for different

slope angles using a safety factor of 1.20. The minimum

solids content required for maintaining a stable slope

can be determined by:

1. Selecting the desired slope angle and landfill height.

2. Finding the required shear strength to wet unit weight

ratio from Figure 7-9.

3. Assuming a wet unit weight of the sludge and computing the shear strength.

4. Finding the solids content for the computed shear

strength from Figure 7-5.

5. Checking the wet unit weight for the obtained solids

content using Equation 7-4:

γ = γw

1 − (1 − 1

Gs

)

Sc(%)

100

(Eq. 7-4)

where

γ = wet unit weight of the sludge

γw = unit weight of water

Gs = specific gravity of the solid phase of sludge

Figure 7-7. Shear strength versus additive level, Alum Sludge

2 (nonaged) (Cornwell et al., 1992).

Fly Ash

Undrained Shear Strength (PSI)

Lime Treated

Soil

Untreated

Curing Time (Days)

Figure 7-8. Undrained sheer strength versus curing time for

untreated and treated Alum Sludge 1 (Cornwell et

al., 1992).

Table 7-16. Undrained Shear Strength of Alum Sludge,

Untreated and Treated (Wang et al., 1992a)

Sludge

Remolded

(psi)

Cured

(psi)

Strength Gain

Ratio

Untreated 0.30 1.70 5.7

Treated with soil 0.42 2.10 5.0

Treated with fly ash 0.82 2.50 3.0

Treated with lime 1.39 3.25 2.3

134
Sc = solids content in percentage

6. Repeating the procedure from step 3 until satisfactory results are achieved.

The shear strength data were also used to analyze the

stability of heavy construction equipment operation on

the landfill. The types of equipment analyzed were

crawler dozer, dump truck, and excavator. The shear

strength and solids content required to support the various types of heavy equipment with a safety factor of 3.0

are summarized in Table 7-17. The minimum shear

strength required to support heavy equipment varies

greatly among the different types of equipment, and the

data were obtained for a safety factor of 3.0. When a

smaller safety factor is used, the required minimum

shear strength is reduced. For example, for a safety

factor of 1.5, the required minimum shear strength to

support the excavator decreases from 19.8 to 9.8

kiloPascals. This value is very close to 10 kiloPascals

of the European standard (Cornwell and Koppers,

1990). It is not known, however, whether the European

standard was developed for the conditions stated. The

minimum shear strength required to support heavy

equipment should be selected with consideration of the

factors that influence equipment stability: equipment

type (weight, dimension, wheel design, and nature of

loading), sludge layer thickness, sludge surface area

and condition (level or sloped), and safety factor, among

other factors.

According to Table 7-17, the minimum solids concentration required to support heavy equipment varies considerably among the three sludges. Thus, the required

minimum solids concentration must be determined for

every sludge, precluding a single value of minimum

solids concentration for general use. The minimum solids concentration for Alum Sludge 1 in Table 7-17 reportedly correlates well with field conditions (Cornwell et al.,

1992). Therefore, the data should be useful in determining minimum solids concentration and other information

pertinent to landfilling of water treatment sludges.

Landfill Slope

9H:1V


8H:1V


7H:1V


6H:1V


5H:1V


4H:1V


3H:1V

Shear Strength/Unit Weight (M)

Landfill Height (M)

Figure 7-9. Landfill height versus shear strength/unit weight for

different slope angles (Wang et al., 1992b).

Table 7-17. Required Shear Strength and Solids Concentration for Hypothetical Monofill Supporting Various Types of Heavy

Equipment (Cornwell et al., 1992)

Ground Pressure

Minimum Shear

Strength Approximate Sludge Concentration (%)

Type (psi) (kN/m2

) (psi) (kN/m2)

Alum

Sludge 1

Alum

Sludge 2

Ferric

Sludge 3

Crawler dozer 4.2 28.7 2.0 13.7 29 22 52

Crawler dozer 6.7 45.8 3.3 22.6 33 25 54

Excavator 5.9 40.4 2.9 19.8 32 23 53

Dump truck (empty) 25 171.0 12.2 83.5 37 28 58

Dump truck (full) 50 342.0 24.4 166.9 41 33 62

135
Chapter 8

Land Application

Land application is an increasingly popular disposal

option in the water supply industry due to the escalating

regulatory and environmental constraints associated

with other disposal options. Land application options for

water treatment plant (WTP) residuals include agricultural use, silvicultural application, application for reclamation to disturbed and marginal lands, and disposal to

dedicated lands (U.S. EPA, 1995b). In addition to the

obvious advantages of waste disposal, this application

process may beneficially modify soil properties while

recycling residual components. Potential disadvantages

of land application of WTP residuals include an increase

in the concentration of metals in the soil and possibly in

ground water; adsorption of soil phosphorus by water

residuals, decreasing the productivity of the soil; excessive application of nitrogen, resulting in the transport of

nitrate to ground water; and possible effects caused by

the application of poorly crystallized solids of aluminum

(Dempsey et al., 1990).

This chapter provides decision-makers with the basic

information they need to evaluate land application as a

disposal option for WTP residuals. A simplified planning

procedure for land application is presented in Figure 8-1.

8.1 Regulatory Requirements

The Water Pollution Control Act of 1972 defined WTP

residuals as industrial waste, and they are listed as a

solid waste under 40 CFR Part 257. If land application

is to be considered a viable option, the residuals cannot

be hazardous. Section 2.7 outlines a method for determining whether the residuals are hazardous. Other than

those listed in 40 CFR Part 257, specific federal regulations do not exist for the water industry, and most state

regulatory agencies use the same criteria for WTP

wastes as they use for other industrial sludges. According to 40 CFR Part 257, criteria in the following areas

must be met: floodplains, endangered species, surface

water, ground water, disease vectors, air emission, and

safety. The regulations also contain application limits for

cadmium (Cd), and polychlorinated biphenyls (PCBs).

The Clean Water Act (40 CFR Part 503), establishes

criteria for land application of sewage sludge solids

generated during the treatment of domestic sewage.

Although these criteria are only designed for sewage

sludge, or biosolids, they may provide general guidelines for use or disposal of WTP residuals. The guidelines contain numerical values for biosolids application

rates that are based on experiments with biosolids, and

risk assessment algorithms developed using the

biosolids data. The criteria should not be used for determining land application rates of WTP residuals.

The risk assessment algorithms that were used as a

technical basis for the Part 503 rule contain data that

were obtained from field studies using biosolids. There

is no way to determine, without similar field studies done

specifically on WTP residuals, whether residuals would

demonstrate the same behavior as that of biosolids.

Therefore, land application of WTP residuals should be

based on the criteria set forth in 40 CFR Part 257, and not

on the criteria developed under Part 503. If WTP residuals are field tested and their behavior relative to

biosolids is known, then it is possible that the risk assessment methodology used in Part 503 could be used to

develop criteria for WTP residual use in land application.

Regulations vary from state to state but typically require

that residuals be tested for total metals concentrations,

nutrients, and pH. The application rate of WTP sludge

may be selected to limit hydraulic loading, nitrogen content, or metal content.

8.2 Environmental Considerations

When land application was originally investigated as a

potential disposal option for WTP residuals, it was

thought that the lack of organic components and the

high metals content in WTP residuals would inhibit the

water industry from accepting this option. Since then, a

number of studies have examined the potential limiting

effects of WTP residuals on soil parameters. Most studies to date have focused on identifying the effects of

WTP residuals on the physical characteristics of soil and

on determining the potential for these wastes to cause

phosphorus deficiency in plants. Current research has

yielded a better understanding of the chemical characteristics of WTP residuals and has produced analytical

techniques to quantitatively predict these effects when

136
applied to land. The conclusions of this research are

described below.

8.2.1 Major Components of Water Treatment

Residuals and Their Impact on Soil

Parameters

A 1990 American Water Works Association Research

Foundation study of land application of WTP residuals

was conducted to establish a scientific basis for land

application of WTP residuals. This study examined data

regarding the chemical composition and fractionation of

WTP residuals, the phytotoxicity of WTP residuals, and

the effects of these residuals on the physical and nutrient status of soil. The study also investigated concentrations of residuals’ major components, and the impact

these components might have on agricultural soil after

land application. Table 8-1 shows the composition of WTP

residuals compared with sewage sludge and agronomic

soils in the United States. The cadmium reading of 5.15

parts per million (ppm) is much higher than is typically

found in WTP residuals. This is particularly true now that

chemical suppliers are more conscious of heavy metals

contaminants in coagulants, the major source of metals

in residuals. Most cadmium levels in residuals are less

than 1 to 2 ppm and are often undetectable.

WTP residuals are predominantly inorganic and consist

of clays, humic substances, and other materials typically

suspended or dissolved in lakes and streams. Typical

WTP residuals contain 3 percent or less organic carbon

by weight and 0.5 percent or less organic nitrogen by

Determine Physical and Chemical Sludge Characteristics


(Chapter 3 and Section 8.2.1)

Review Applicable Regulations and


Guidelines for Land Application


Federal, State, and Local (Chapter 2 and Section 8.1)

Compare Sludge Characteristics With


Regulatory Requirements and


Choose Land Application Option

Select Application Site Based on Acreage Required for


Sludge Application, Costs, and Land Area Available


(Section 8.3)

Determine Sludge Transportation and Application Systems


Based on Operational Considerations and Costs


(Section 8.4.1 and Section 8.4.3)

Determine Storage Requirements and Rates of Application


(Section 8.4.1)

Develop Soil and Ground-water Monitoring Programs


(Section 8.4.4)

Agricultural and


Silvicultural Application


(Sections 8.3.1 and 8.3.2)

Land Reclamation


(Section 8.3.3)

Dedicated


Land Disposal


(Section 8.3.4)

Begin Public Participation


(Section 8.4.2)

Figure 8-1. Simplified planning procedure for land application of WTP residuals.

137
weight. Both values are representative of agricultural

soils but are much lower than those found in sewage

sludge. WTP residuals are similar to fine-textured soil.

Elliott et al. (1990a) reported no coliform organisms in

alum or ferric WTP residuals. This absence may be

attributed to clean raw water, disinfection processes

during treatment, and/or lagoon storage of residuals

over long periods. In a waste stream recycling study

(Cornwell and Lee, 1994), however, Giardia and Cryptosporidium were found to concentrate in the sedimentation basin residuals. The study did not examine the

issues of the viability and persistence of these pathogens, nor did it discuss the survival of pathogens in WTP

residuals over time.

WTP residuals largely consist of materials that are removed from water before its distribution. Coagulant addition can enhance turbidity removal and contribute to

the quantity of residuals. Depending on the quality of the

intake water, the final residuals can be clay-like or sandlike in texture, which may become important in choosing

the final use. WTP residuals by their nature are usually

low value fertilizer, as can be seen in Table 8-2. The

concentration of aluminum or iron in WTP residuals is

generally 5 to 15 percent of dry solids. These concentrations stem from the use of coagulants in the water

treatment process and may be higher when a specific

coagulant is overdosed. The aluminum content of WTP

residuals treated with alum is similar to aluminum concentrations in soils. The speciation of aluminum or iron

in WTP residuals, however, is different from that occurring in native soils. While soils contain significant concentrations of aluminum in the form of aluminosilicates,

WTP residuals contain aluminum (or iron) in amorphous

hydrous oxide forms, which exhibit greater reactivity

than the corresponding soils materials. The various aluminum forms, however, quickly convert to the stable

aluminum oxides away from aluminum hydroxides, and

do not exhibit reactivity.

Moderate applications of WTP residuals may improve

the physical condition, or tilth, of soils by flocculating

colloidal particles, thereby promoting soil aggregation

via reactions that are analogous to those occurring during water treatment. The hydrous metal oxides contained in WTP residuals are strong adsorbents of trace

metals and phosphorus. In terms of trace metals, this is

beneficial to soils and results in a decreased and buffered concentration of free trace metals. Elliott et al.

(1990b) concluded that in the case of soils treated with

WTP residuals, soluble aluminum usually decreases,

and the level of the dithionite-citrate-bicarbonate (DCB)1

extractable aluminum residuals is typically elevated over

original levels but remains within the range normally

observed for native soils (see Table 8-1). The amorphous hydrous oxides in alum-treated WTP residuals

may also benefit coarse soils by increasing the cation

exchange capacity, particularly at the neutral or slightly

Table 8-2. Agronomic Components in WTP Residualsa

Parameter Range of Valuesb

Total solids (%) 8.14–81

Volatile solids (%) 9.32–29.09

Conductivity (μmho/cm) 563.8 (530.2)

pH (S.U.) 6.98–8.82

Total Kjeldahl nitrogen (%) 0.495 (0.256)

Organic nitrogen (%) 0.752 (0.399)

Ammonia–nitrogen (%) 0.016 (0.016)

Nitrate–nitrite (%) 0.003 (0.003)

Total phosphorus (%) 0.226 (0.248)

Total potassium (%) 0.225 (0.317)

a Data are presented as a range or as arithmetic mean and standard

deviation ( ). These data should be used with caution. Individual

residuals may vary significantly from those used in this table. b Survey of 12 land-applied WTP residuals, including 10 alum and 2

ferric (Brobst, 1994).

Table 8-1. Composition of WTP Residuals Compared With

Sewage Sludge and Agronomic Soilsa

Constituent WTP Residualsb Sewage

Sludgec Agronomic Soil

Aluminum 60,111.0 (52,095) 72,000d

Arsenic 14.7 (14.7) 9.93 (18.84) 7.2d

Barium 122.5 (127.52) 580d

Boron 40.7 (61.3) 33d

Cadmium 5.15 (11.67) 6.94 (11.76) 0.265 (0.253)e

Calcium 20,815.0 (33,108) 24,000d

Chromium 49.5 (56.5) 118.6 (339.2) 54d

Copper 270.2 (326.1) 741.2 (961.8) 29.6 (40.6)e

Iron 52,751.0 (63,642) 26,000d

Lead 79.9 (100.2) 134.4 (197.8) 12.3 (7.5)e

Manganese 385.4 (398.5) 9,000d

Mercury 1.5 (2.5) 5.2 (15.5) 0.09d

Molybdenum 9.1 (12.9) 9.2 (16.6) 0.97d

Nickel 37.8 (53.8) 42.7 (94.8) 23.9 (28.1)e 

Selenium 2.8 (1.5) 5.2 (7.3) 0.39d

Strontium 84.3 (60.6) 240d

Vanadium 35.4 (17.4) 80d

Zinc 1,047.4 (3,036.1) 1,201.9 (1,554.4) 56.5 (37.2)e

a Data are arithmetic mean and standard deviation ( ) presented as

mg/kg dry weight. b Survey of 12 land-applied WTP residuals, including 10 alum and 2

ferric (Brobst, 1994). c

 U.S. EPA, 1990a.

d Shacklette and Boerngen, 1984.

e Holmgren et al., 1993.

1 DCB is an extracting solution that removes the more reactive fraction

of the total iron and aluminum present in a substance.

138
alkaline soil pH that can increase the nutrient holding

capacity.

8.2.2 Effects of Trace Metals Concentrations

on Soil Properties

Trace metals in most WTP residuals originate from coagulant chemicals and or the water source. These metals are strongly adsorbed to the aluminum hydroxide or

ferric hydroxide precipitates formed during the coagulation process. Concentrations of trace metals will likely

determine the lifetime application of water treatment

residuals (metric tons per hectare, dry weight) on agricultural soils.

Elliott et al. (1990a) assessed potential metal mobility

under field conditions using a five-step fractionation procedure. Most heavy metals in WTP residuals are bound

in forms not readily released into solution due to their

strong adsorption and coprecipitation by aluminum (Al)

and iron (Fe) hydroxides freshly formed in the coagulation treatment step. From 76 to 87 percent of the chromium (Cr), copper (Cu), nickel (Ni), lead (Pb), and zinc

(Zn) examined were bound within an oxide or silicate

matrix. Less than 6 percent of the total concentration of

each of these metals was in the exchangeable fraction,

which is considered to represent the easily available

metal pool. If soil conditions should become acidic (pH

less than 5), approximately 25 percent of the residuals

cadmium could become mobile.

Greenhouse pot studies showed that with the more

acidic aluminum-based WTP residuals, application rates

of 20 and 25 g/kg increased the cadmium concentration

of plant tissue to greater than 2 mg/kg (Heil and Barbarick, 1989). Figure 8-2 illustrates the partitioning of

trace metals in WTP residuals. For example, most chromium is released upon dissolution of Al(OH)3(s) and

Fe(OH)3(s) (i.e., the iron-manganese oxide-bound fraction), whereas a very small portion of chromium is released in the acid-soluble fraction. Similar results can

be obtained for other trace metals. Soil pH conditions

are critical in preventing the dissolution of Al(OH)3(s) or

other hydrous metal oxides; if nonacidic soil conditions

exist, trace metals will be tightly bound, rendering them

immobile and unavailable for plant uptake. Therefore, in

the case of moderate WTP residuals application rates

(20 dry mt/ha) and properly managed soils (pH greater

than 6.0), movement of metals into ground water or into

plant tissues can be minimized.

8.2.3 Impact of Water Treatment Residuals

on the Availability of Phosphorus in

Agricultural Soils

Because phosphate deficiency greatly affects agricultural soil fertility, much research has been conducted

into the decreased availability of phosphorus in soils

following the addition of WTP residuals. This decrease

in available phosphorus occurs because of the large

amounts of aluminum and iron hydroxide solids present

in WTP residuals. These hydroxides are strong adsorbents of inorganic phosphorus and contain only low

concentrations of phosphorus. This phosphorus-limiting

characteristic of WTP residuals is illustrated in Figure 8-3.

Greenhouse studies found phosphorus deficiencies in

sorghum-sudangrass were possible when application

rates of aluminum- or iron-based WTP residuals exceeded 10 g/kg soil or approximately 22.5 mt/ha (Heil,

1988; Heil and Barbarick, 1989). The study also found

that both alum- and iron-based WTP residuals decrease

the availability of phosphorus but can effectively increase iron availability in iron-deficient soil. WTP residuals applied at a rate of 60 g/kg or approximately 135

mt/ha, produced phosphorus deficiencies in tomato

plants (Elliott and Singer, 1988).

This reduction in plant-available phosphorus is qualitatively different from the potential toxicity problem that

may be associated with land application of WTP residuals. The depletion of plant-available phosphorus necessitates the addition of supplemental phosphorus

100


90


80


70


60


50


40


30


20


10


0

Cd Cu Pb

Cr Nl Zn

Exchangable

Organically Bound

Acid Soluable

Fe-Mn Oxide Bound

Residual

% Total Metals

Definitions:

1. Exchangeable: 16 mL 1M MgCl2 (pH 7.0); shake time, 1 hr.

2. Dilute acid extractable: 16 mL 1M NaOAc adjusted to pH 5.0

with HOAc; shake time, 5 hrs.

3. Fe-Mn oxide bound: 40 mL of 0.175M (NH4)2C2O4 and 0.1M

H2C2O4; shake time, 4 hrs.

4. Organically bound: 40 mL of 0.1M Na4P2O7; shake time, 24 hrs.

5. Residuals: Dry 0.1 g of material remaining after 4/. Add 4 mL

HNO3, 1 mL HClO4 and 6 mL of HF, heat at 140°C for 3.5 hrs.

Add 5.0 g Boric acid and dilute to 100 mL volume.

Figure 8-2. Partitioning of trace metals in WTP residuals (after

Elliott et al., 1990a).

139
fertilizer to obtain optimal plant yields. This is a management problem rather than a public health issue; therefore, the problem is easy to correct with either adjusted

management practices or additional phosphorus fertilizer.

Soluble phosphorus levels are commonly determined by

the equilibrium phosphorus concentration (EPC) test

(White, 1981). The EPC is the concentration of phosphorus at which no adsorption or desorption of phosphorus occurs. The phosphate buffer capacity (PBC)

provides information about the response of a soil to

fertilization, specifically whether the added phosphorus

will be adsorbed by the soil or will remain in the soil

solution and be available for plant uptake. Elliott et al.

(1990b) used both EPC and PBC techniques to investigate the incremental amount of phosphorus fertilizer

that should be applied to soils treated with WTP residuals to obtain good plant growth.

The typical range of EPC values for optimal production

is 50 to 200 μg/L EPC. Between 50 and 100 of the triple

superphosphate (TSP) fertilizer used in this study is

conventionally added to agricultural soils to maintain

crop production (White, 1981). Farmers usually apply

nitrogen and potassium fertilizers to agricultural soils in

addition to various pesticides. Therefore, the cost of

restoring the EPC after residuals application should be a

fraction of the entire cost of preparing a soil for planting.

Application techniques may be used to mitigate the

impact of WTP residuals on the availability of phosphorus in agricultural soils. WTP residuals can be applied

to the top soil layers or injected below the surface,

depending on the growth stage of specific crops. Seedlings generally have higher phosphorus requirements

than mature plants do; therefore, deep injection (greater

than 18 cm) may be preferable at this growth stage.

When residuals are applied to the soil surface, the resulting drying action decreases the surface area of the

aluminum and iron hydroxides, and organic matter in the

surface litter will localize the adsorption sites for phosphorus. This may in fact be beneficial in areas of high

naturally occurring phosphates, and WTP residuals may

be able to reduce phosphorus loadings to surface waters by adsorbing the phosphorus. In addition, in areas

in which stormwater runoff carries excess phosphorus

to lakes and streams, WTP residuals could be used to

bind excess phosphorus and improve the runoff water

quality.

Crops may be selected that tend to use labile phosphorus rather than just soluble phosphorus. Large quantities

of phosphorus fertilizer are applied in close proximity to

seeds to accommodate the phosphorus requirements of

seedlings. Sewage sludge contains significant amounts

of phosphorus that can add to total soil phosphorus.

Several Colorado facilities have studied the coapplication of both sewage sludge and WTP residuals to mitigate the effects of phosphorus adsorption by WTP

residuals. The experiments have met with varying degrees of success in the short evaluation periods, however, they have shown that the assumption of

phosphorus being bound on a one-to-one molar basis,

as required by state regulations, may be conservative.

In the case of excessively fertilized soils, phosphorus

adsorption by WTP residuals may be agriculturally advantageous. WTP residuals applied to high phosphorus

soils can produce a lower EPC for optimal crop growth.

8.2.4 Effects of WTP Residuals on Soil

Physical Properties

Research has been conducted to determine whether

application of WTP residuals conditions agricultural soils

by increasing the soil organic matter content. Soil conditioning is treatment that modifies a soil’s physical properties for the improvement of crop growth (Chang et al.,

1983). The organic matter content of dry WTP residuals

varies depending on the original raw water source and

the treatment methods employed at individual WTPs.

Estimates of percent organic matter in WTP residuals

range from 14.4 percent (Lin, 1988) to 25 to 35 percent

(Bugbee and Frink, 1985). These values seem high and

should be used with caution. Appropriate methods for

determining organic matter in WTP residuals should be

used. Current research in this field (Elliott et al., 1990a)

determined that the average organic matter content of

seven test residuals was 3 percent, with an average loss

on ignition (LOI) value of 33 percent. Thus, the soil

conditioning effects of organic matter in WTP residuals

are probably quite small and may not be measurable

under field conditions.

5


4


3


2


1

0 10 20


 5 15 25

Average Alum


Average Iron

WTP Residuals Application Rate (g/kg)

Plant Tissue P (g/kg)

Figure 8-3. Average (of three cuttings) phosphorus concentration in sorghum-sudangrass grown in Colby soil

(after Heil and Barbarick, 1989).

140
At a land application site in Colorado with cumulative

application rates of up to 179 mt/ha (80 tons/acre) (by

1984) researchers theorized that the cumulative loadings of quartz and aluminum hydroxides exhibited some

potential to change soil texture in fields near an industrial water treatment facility (King et al., 1988). The

annual application rate was approximately 45 mt/ha (20

tons/acre). Soils on the fields amended with alum residuals have shown a potential for dusting and more

rapid drying than the nonamended fields. Application

rates were lowered to control this situation.

8.3 Land Application Options

8.3.1 Agricultural Land

Agricultural use of WTP residuals is the beneficial application or disposal of residuals on agronomic lands. Successful implementation of this disposal option requires

an understanding of the effects of WTP residuals on soil

fertility and physical properties (discussed in the previous section). The physical characteristics of soil that

determine whether it can support vegetative growth include cohesion, aggregation, strength, and texture.

These parameters directly affect the hydraulic properties of a soil, such as moisture-holding capacity, infiltration, permeability, and drainage. Any adverse impact on

these hydraulic soil characteristics from land-applied

WTP residuals can effect crop growth and ultimately

degrade ground water quality (Elliott et al., 1990b).

Concentrations of the WTP residuals listed in Table 8-1

tend to be elevated compared with the concentration of

trace metals described in Chapter 3. Many of the sampled WTPs in Colorado treat surface waters affected by

past mining activities. The length of time a site may be

used depends on the quality of the WTP residuals applied. Two basic types of residuals are generated at a

WTP—lime softening residuals and coagulant (alum or

ferric) residuals. These residuals are discussed below.

8.3.1.1 Lime Softening Residuals

Addition of lime to agricultural soil is a common practice

in areas where the soil pH is too low for optimal plant

growth; lime modifies the balance between acidity and

alkalinity in the soil. Soil pH should be maintained at 6.5

or above to minimize crop uptake of metals (U.S. EPA,

1983). In addition to providing a desirable pH for plant

growth, lime residuals can be substituted for agricultural

limestone and offer calcium carbonate equivalence

(CCE) or neutralization effectiveness. Lime residuals

can be dewatered to higher solids content than alum

residuals can; therefore, the potential for unique disposal or co-disposal opportunities is greater than for

other residuals. Studies show the neutralizing power of

lime residuals is equal or superior to that of agricultural

limestone (Che et al., 1988). Lime residuals increase the

porosity of tight soils, rendering the soils more workable

for agricultural purposes.

The best way to estimate the liming requirement of a soil

is to titrate individual samples of the soil with a standard

base. The result is the quantity of base required to raise

the pH of the soil to a specified pH. The theoretical

chemical process is illustrated below:

CaCO3 + H2O = Ca2+ + HCO3 − + OH−

Ten mmoles of a base per kilogram consumed during

titration is equivalent to 4.5 metric tons of pure CaCO3

per hectare at a 30-cm plough layer (ha-30 cm) of the

sample soil.

Field liming results may differ considerably from these

experimental results, depending on the mixing of lime

residuals with the soil, the solubility of the liming agent,

and the soil sampling procedures. Use of a conversion

factor generally corrects this difference. The U.S. Department of Agriculture’s Extension Service may be consulted for the appropriate conversion factor for a given area.

Lime residuals can be applied as a liquid or dewatered

and applied as a cake using an agricultural manure

spreader. The application rate for lime residuals must be

calculated on a dry weight basis, then converted to the

equivalent wet weight for application to the farm field.

8.3.1.2 Alum Residuals

Alum residuals when dry generally have the consistency

of very fine soils. Bugbee and Frink (1985) indicate that

alum residuals improve the physical characteristics of

soil media but inhibit plant growth by adsorbing phosphorus. As stated earlier, this reduction in phosphorus

available to plants is different from the toxicity problem

that may be associated with land application of WTP

residuals. To obtain optimal plant yields when land applying alum residuals, phosphorus fertilizer must be added.

As previously mentioned, soil conditioning effects from

organic matter in WTP residuals are probably quite

small and may not be measurable under field conditions.

Management practices such as lowering application

rates can control the dusting and rapid drying seen in

fields amended with alum residuals. Application of alum

residuals at rates that are not considered excessive (20

mt/ha) does not cause environmental degradation. Alum

residuals are applied as a liquid unless the WTP has

dewatering capabilities. Liquid alum residuals can be

applied with a liquid manure spreader, like lime residuals, or with conventional irrigation equipment.

Example 1: Co-disposal With Sewage Sludge

Colorado in the late 1980s developed a WTP residuals

land application policy based on research completed in

northeastern Colorado. This approach assumed that

aluminum and iron would bind with the phosphorus in

141
the WTP residuals on a one-to-one basis. Because of

this potential phosphorus binding effect, Colorado developed a conservative method of estimating the fraction of WTP residuals that can be coapplied with sewage

sludge. That method is described in Equation 8-1.

WTP residual

fraction =

P(SS) − 1.15 Al(SS) −

0.55 Fe(SR)

[1.15 Al(WTR) − 1.15 Al(SR)] +

[0.55 Fe(WTR) − 0.55 Fe(SR)]

(Eq. 8-1)

where

P(SS) = phosphorus content (mg/kg) in sewage

sludge

Al(SS) = aluminum content (mg/kg) in sewage

sludge

Fe(SS) = iron content (mg/kg) in sewage sludge

Al(WTR) = aluminum content (mg/kg) in WTP

residuals

Fe(WTR) = iron content (mg/kg) in WTP residuals

For example, if

P(SS) = 25,000 mg/kg (average value from U.S.

EPA, 1995)

Al(SS) = 8,000 mg/kg (typical value)

Fe(SS) = 15,000 mg/kg (typical value)

Al(WTR) = 60,111 mg/kg (average value from Table

8-1)

Fe(WTR) = 52,751 mg/kg (average value from Table

8-1)

then, the WTP residuals fraction = 

25,000 mg⁄kg P − 1.15 (8,000 mg⁄kg Al) −

0.55 (15,000 mg⁄kg Fe)

[1.15 (60,111 mg⁄kg Al) − 1.15 (8,000 mg⁄kg Al)] +

 [0.55 (52,751 mg⁄kg Fe) − 0.55 (15,000 mg⁄kg Fe)]

= 0.193

The values used in this example are only for demonstration purposes. Because each WTP residual and sewage

sludge has different qualities, the most current data from

a given treatment plant should be used. It should be

noted that Colorado is no longer using this method and

has been evaluating land application requests on a

case-by-case basis.

The result of this equation and agronomic information

for the sewage sludge can be used to determine a land

application rate for residuals. If the nitrogen in a WTP

residual is higher than expected (higher than in Table

8-2), the nitrogen content must also be used in calculating the agronomic rate. The agronomic rate of sewage

sludge is the amount needed to supply the recommended amount of nitrogen (N) for a particular crop

without allowing excess nitrogen to migrate below the

root zone and into the ground water. According to 40

CFR Part 503.14.d, “Bulk sewage sludge shall not be

applied at rates above agronomic rates, with the exception of reclamation projects and when authorized by the

permitting authority.”

When calculating the agronomic rate, note that soil data

are required to determine the amount of nitrogen

needed for a specific crop. Therefore, for dry land farming, deep soil monitoring should be conducted once

every 5 years, with samples collected at 1-foot intervals

for a total of 5 feet. For irrigated farm land, samples

should be collected within the plow zone on an annual

basis. If soil data are not available, residual nitrogen

from previous sewage sludge applications can be determined using the mineralization rate of organic nitrogen.

Annual soil tests, however, provide a more accurate

estimate of the nitrogen content of the soil. Because an

individual soil sample may not provide an accurate determination for an entire land area, multiple soil samples

should be taken from different locations within the field.

Example 2: Land Application With Supplemental

Phosphorus

Fertilizer recommendations are specific to crop, climate,

soil type, and yield goals. Therefore, this option is discussed only generally.

Generally, the phosphorus (P) soil test rates soils as low,

medium, high, and very high with respect to available

phosphorus. Thus, the soil test is interpretive rather than

purely analytical. For soils with low available phosphorus, as indicated by the test, good agricultural practice

is to add significantly more phosphorus than the crop

needs. This extra phosphorus is added to compensate

for the phosphorus-depleting effects of land-applied

alum/iron residuals in addition to the intrinsically low

phosphorus availability. For soils rated as medium,

slightly more phosphorus should be added than the crop

demands. Soils rated as high in phosphorus need none

added; however, enough phosphorus fertilizer should be

added to compensate for any phosphorus depletion occurring due to the presence of alum/iron residuals in the soil.

For example, assume that the phosphorus demand for

WTP residuals applied at 1 mt/ha is equivalent to 98.13

kg/ha (196.26 lb/acre). Also, the soil test indicated that

soil phosphorus demand is 15 mg/kg. Using Table 8-3’s

figures for a wheat crop:

• Phosphorus demand from the alum/iron residuals =

98.1 kg/ha.

• Phosphorus required to satisfy the crop need (from

the table) = 19.5 kg/ha.

Therefore,

• Estimated soil demand for phosphorus = 5 kg/ha.

The total supplemental phosphorus required for this

example is 132.6 kg/ha (265 lb/acre), with the alum/iron

142
residuals applied at 1 mt/ha. Again, the assumptions

made here are for illustration purposes only; the WTP

must modify this information to reflect local conditions.

8.3.2 Silviculture

Silviculture, another agricultural practice, is the establishment, growth, reproduction, and care of forest trees

as a crop. Land application of both wastewater and WTP

residuals is relatively uncommon in silviculture compared with application to agricultural lands. Bugbee and

Frink (1985) indicated the following effects of residuals

application to silvicultural use: no significant effects on

tree growth, nutrient levels, or the appearance of the

forest floor; slightly diminished uptake of phosphorus;

and increased soil pH for trees treated with WTP residuals. Grabarek and Krug (1987) concluded that limited

application of WTP residuals to a healthy forested area

does not upset the phosphate cycle to the point of

affecting forest growth patterns.

The effects of long-term (30 months) application of alum

residuals were evaluated by Geertsema et al. (1994).

Alum residuals were applied to experimental plots with

two to four replicates of each application rate: 0, 36, and

52 mt/ha (0, 16, and 23 dry tons per acre). The study

found no statistically significant differences in soil characteristics, ground-water characteristics, or loblolly pine

growth analyses between the unamended and the residuals-amended plots. Geertsema et al. also found no

significant metal migration through the soil profile. At the

application rate used in the study, no problems were

observed in available phosphorus as evidenced by the

similar pine tree growth in the treated and untreated

plots. The results of the study did not indicate any environmental problems; therefore, the City of Newport

News has implemented a full-scale project.

8.3.3 Land Reclamation

Sewage sludge is more commonly used than WTP residuals to reclaim surface-mined areas or other disturbed lands and to establish vegetative growth and/or

restore or enhance soil productivity. Disturbed lands

include highway construction sites, overgrazed rangelands, and other construction sites. In combination

with other fertilizers, however, WTP residuals may benefit reclamation efforts. WTP residuals can be used to

treat a particular or site-specific concern. For example,

lime residuals can control soil pH just as they do in

agriculture, but in these cases the pH adjustment may

be more critical because mine soils can be very low in

pH. In addition, alum residuals can control runoff of

excess phosphate into surface waters. Care must be

taken to ensure that the site is suitable for use of WTP

residuals and that controls (best management practices)

are in place to protect public health and the environment.

Application rates may vary from 5 mt/ha on arid rangeland to 450 mt/ha or higher on a mineland reclamation

site. On a mineland reclamation site, WTP residuals

may actually be used as a topsoil replacement; thus, the

application rate may be as high as 450 mt/ha.

8.3.4 Dedicated Land Disposal

Dedicated land disposal (DLD) of WTP residuals is generally the alternative chosen by those operations that

either generate residuals continuously or have quantity

or quality concerns. DLDs are designed to treat and

dispose of large quantities of residuals through soil microorganisms, sunlight, and/or oxidation. In addition, the

soils are used to bind or fix the metals, thus making them

unavailable.

A distinct advantage to using a DLD is that, for the most

part, the WTP controls the application times, rather than

other factors (e.g., crop growth cycles). DLD application

rates are generally much higher than those of other land

application operations, ranging from 20 mt/ha to greater

than 200 mt/ha. This increased application rate can greatly

decrease the amount of land required for the WTP.

The tradeoff between DLDs and other land application

options is that stricter requirements generally apply to

the DLDs. Increased application rates require that the

site be carefully designed, managed, and monitored.

The site must be designed to contain any of the residual

constituents that might threaten public health or the

environment. The site must take more care with surface

runoff controls and the generation of contaminated

leachate or infiltration. Generally, surface water controls

include physical structures such as dikes, ditches, and

lagoons. Control of leachate that is generated may require collection and treatment prior to discharge or disposal. Discharge of the treated leachate may require a

National Pollutant Discharge Elimination System (NPDES)

permit and/or additional state permits.

Table 8-3. Phosphorus Recommendations for Several

Agronomic Cropsa (Foth and Ellis, 1988)

Phosphorus Applied kg/ha (lbs/acre)

Soil P Test

mg/kg Corn Wheat Soybeans Potatoes

15 15 (30) 19.5 (39) 11 (22) 38 (76)

30 6.5 (13) 11 (22) 0 35 (70)

50 0 0 0 30.5 (61)

80 0 0 0 24 (48)

175 0 0 0 0

a Yield goals assumed: corn, 140 bushels/acre; wheat, 70 bushels/

acre; soybeans, 50 bushels/acre; potatoes, 400 cwt/acre.

143
8.3.5 Other Use Options for WTP Residuals

Other applications for WTP residuals include use at turf

farms, parks and recreational areas, highway rights of

way, airports, and construction sites. Most of these projects utilize WTP residuals as a soil replacement or soil

amendment. For example, in turf farming, each time a

crop is harvested, a significant amount of soil is removed. WTP residuals can be used as a partial replacement. The other options also typically require imported

soil, and in areas where topsoil is scarce residuals may

be used as a partial or total replacement. As with all land

application projects, however, the quality of the residuals must be matched with the needs of the user.

8.4 Operational Considerations in Land

Application

8.4.1 Application Procedures

Once land application is selected as the preferred disposal option, the next step is to evaluate the land application site, the costs associated with that site, and the

potential social and environmental impacts on the site.

In addition, the costs of dewatering and hauling must be

reviewed because the higher the percentage of solids,

the less volume of residual must be hauled.

The land area requirement depends on the application

rate. In general, application rates are determined using

agricultural methods, as with fertilizer. In the case of

WTP residuals, this fertilizer recommendation is at least

two fold; the nutrient requirements for the crop are

based on the expected yield and the available nutrients

in the soil. The first nutrient to consider is nitrogen. WTP

residuals tend to be very low in this nutrient. A WTP

should always check the nitrogen content of the residuals to be land applied. Basing the application rate on

nitrogen alone, however, may cause problems with the

crop yield because aluminum and/or iron hydroxide solids present in WTP residuals are strong adsorbents of

inorganic phosphorus. The low concentrations of phosphorus in WTP residuals may further diminish total

phosphorus in soils, restricting plant growth.

As previously noted, concentrations of trace metals will

likely determine the lifetime application amount of WTP

residuals (metric tons per hectare, dry weight) on agricultural soils. Most metals in WTP residuals are bound

in forms not readily released into solution due to their

strong adsorption and co-precipitation by aluminum and

iron hydroxides freshly formed in the coagulation treatment step.

Buffer zones between residuals applications, such as

surface water, drinking water wells, drainage ditches,

property lines, residences, schools, playgrounds, airports, and public roadways, play an important role in a

successful WTP residual program. Identification of sitespecific buffer zones may be critical to the land application program. In addition, many states have buffer zone

regulations that must be followed. Guidelines for buffer

zone delineation can be found in the following EPA

publications:

• Process Design Manual Land Application of Municipal Sludge (U.S. EPA, 1995b).

• Biosolids Management Handbook for Small Municipal

Wastewater Treatment Plants, U.S. EPA Regions 7,

8, and 10 (U.S. EPA, 1995a).

Land application may be feasible only at certain times

of the year, depending on the moisture conditions of the

soil and weather conditions. Another, often overlooked

factor is the local cropping cycle. Farmers will want to

restrict application to certain times of the year, which will

require additional or supplemental storage of WTP residuals. This timing and storage must be factored into

the land application program.

Land application of WTP residuals can be accomplished

through injection, incorporation, or surface application.

Surface application requires the least additional equipment. Land application equipment is widely available,

often from farm implement dealers.

8.4.2 Public Participation

Involving area residents may seem an obvious step but

is often not considered. Those responsible for the use

or disposal of WTP residuals must involve the community before starting the project. Holding public hearings

is only a small part of a broad-based participatory process. Often the public can help make necessary decisions. Given information and allowed to participate, the

public often becomes an advocate for the WTP. Target

community leaders, place information in libraries, and

talk to people. Community leaders include elected officials, church pastors, community activists, and others.

Consensus on the best solution is gained only through

the joint effort of all involved: regulators, contractors,

and the public.

8.4.3 Transportation

Transportation of WTP residuals represents a major

cost and should be carefully examined. The first consideration is the type of WTP residuals being handled and

transported, liquid or cake. The lower the percentage of

solids, the more liquid the residuals, and the greater the

volume of residuals that must be handled and transported.

Liquid can be transported either by pipeline or by truck.

Pipeline offers the convenience of transporting residuals

regardless of weather or other external factors. The

distinct disadvantage, however, is that pipelines generally go to one location and limit the options for land

application. To overcome this disadvantage, the WTP

144
can maintain a large tract of land, enough to handle the

residuals generated. Truck transportation offers WTPs

more options for land application; therefore, most land

application programs use this transportation method.

Before various residuals dewatering and disposal alternatives can be compared, the costs of loading, dewatering, hauling, and disposal must be evaluated.

8.4.4 Monitoring

8.4.4.1 Residuals Sampling and Analysis

Either grab samples or composites can be used to

analyze WTP residuals. WTP residuals range in consistency from liquid to solid; sampling techniques differ for

each residual consistency. The U.S. Environmental Protection Agency (EPA) has developed guidance for sampling methods for sewage sludge that both explain (U.S.

EPA, 1989) and show (U.S. EPA, 1993c) residuals sampling. The goal of all sampling is to obtain a representative sample, which is easiest to accomplish with a

liquid being pumped.

One method to obtain a representative sample from a

residual pile was developed in the early 1900s (first

referenced in 1902) in the mining district of England.

Many samples are taken from different depths and at

different locations in the pile. A cone is built of the

subsamples, and the cone is then flattened and quartered, thus the term “quartering.” Opposite quarters are

re-coned and quartered again. This process continues

until the required sample volume is reached.

Analytical methods for metals are found in the latest

edition of Test Methods for Evaluating Solid Waste (U.S.

EPA, 1986b). A common analytical error is for labs to

conduct the metals analysis using a method developed

for water and wastewater. In addition, the sample preparation method is as important for metals as the correct

analytical method. All metals analysis samples should

be prepared using SW-846 method 3050. With this

preparation method, it is important to use equivalent to

1 gram dry weight of solids for the digestion. The results

should be reported as dry weight, which requires the

analysis of percent total solids so that dry weight can be

calculated.

8.4.4.2 Soil Sampling and Analysis

Soil sampling is essential for accurate fertilizer recommendations. The recommendation may be used to calculate the quantity of WTP residuals to apply. With alum

and ferric residuals, however, supplemental phosphorus

fertilizer may be needed.

Total metals in the soil should be analyzed using the

method described in Methods of Soil Sampling and

Analysis (Page et al., 1982). Total metals should be

analyzed rather than plant-available levels; methods

vary depending on locale, and the scientific community

still has concerns about the comparability of results.

Local land grant universities should be consulted for

more information. As a basic guideline, however, one set

of composite samples should represent no more than 16

hectares.

When sampling an area of a field, samples should be

taken from areas of similar soil characteristics, such as

color, slope, and texture. Areas with different characteristics may require additional samples. A minimum of

20 subsamples for compositing should be taken from the

sampling area, regardless of size. The sample should

be taken to a depth approximately equal to the plow

depth, discarding the surface litter of each sample. Samples can then be air dried and sent to the laboratory for

analysis. Typically, the cost of metals analysis of WTP

residuals ranges from $100 to $250 per sample. Other

analysis, such as TCLP, may cost as much as $1,500.

Soil analyses are similar in cost. The local land grant

university may be contacted for both analysis and fertilizer recommendations.

8.4.4.3 Other Analyses

Other sampling techniques and chemical analyses may

be required to assess a specific site or residual, such as

plant tissue sampling and monitoring, ground-water

monitoring, or deep soil monitoring. Deep soil monitoring samples soils to a specific depth; in EPA Region 8,

depth is 5 feet. Each 1-foot increment is analyzed to

assess migration of pollutants of concern, thus a facility

can use deep soil monitoring instead of ground-water

monitoring.

145
Chapter 9

Brine Waste Disposal

Currently, reverse osmosis (RO), electrodialysis (ED) or

electrodialysis reversal (EDR), nanofiltration (NF), ultrafiltration (UF), and microfiltration (MF) are the only

membrane processes being used or considered for use

in treating public water supplies. All of these processes

produce residuals waste streams, but not all generate

waste streams that could be classified as brine—that is,

as having a concentration of dissolved salt higher than

that of ordinary seawater (Lapedes, 1978; Ingram,

1969). Only seawater RO, ED or EDR, and brackish

water RO can produce brine waste streams. All other

residuals waste streams from membrane processes are

concentrates rather than brine. In fact, the American

Society for Testing and Materials (ASTM) publishes a

standard for definitions of water-related terms (ASTM,

1982) that interprets the residuals portion of an aqueous

solution applied to a membrane as a concentrate. This

chapter focuses primarily on the disposal of brine waste,

with information and background on concentrates supplied as appropriate.

9.1 Background Information

9.1.1 Amount of Concentrate Generated and

Disposal Methods

The U.S. Office of Technology Assessment (OTA) evaluates waste concentrate generation from membrane

processes in terms of percent recovery of feed water

and percent disposal as waste concentrate (Table 9-1).

Conventional methods of concentrate disposal involve

disposal to surface bodies of water, spray irrigation combined with another dilution stream, deep well injection,

drainfields, boreholes, or wastewater collection systems. These methods are usually the most cost-effective. Nonconventional methods are generally considered

cost-prohibitive and are usually associated with a zero

discharge scenario. These include use of evaporation

and crystallization technologies and other ancillary

equipment to concentrate the waste stream into a cake.

Other nonconventional methods include evaporation by

solar ponds or solar distillation.

It should be noted that water treatment sometimes involves ion exchange, a separation process in which

water is passed through a column containing an ion

exchange resin or medium, often a cationic or anionic

synthetic polymer. As the water passes through, the

resins exchange their ions with those in the water according to the preference of the exchange sites for

specific ions and the concentrations of these ions. In this

way, ions from the water adsorb onto the resin. When

the resin approaches its capacity for taking up ions, the

ion exchange process is taken off line and the column

regenerated by exposing the resin to a regenerant solution that, through ion exchange, converts the resin back

to its initial form. The off-line steps include backwashing,

regeneration, and rinsing. After regeneration, the ion

exchange column is returned to service.

9.1.2 Constraints and Concerns

Concentrate disposal methods must be evaluated in

light of geographic, environmental, and regulatory impacts. Geographic constraints generally have to do with

limitations in local geology, hydrology, weather, and land

availability. Environmental concerns such as impacts on

wetlands, flora and fauna, and surface or ground-water

pollution limit the use of some disposal methods in some

locations.

At least eight states have established regulations to

protect water resources from further degradation from

brine or concentrate disposal. Section 305(b) of the

Clean Water Act requires states to report to the U.S.

Environmental Protection Agency (EPA) the extent to

which their waters are meeting the goal of the Act and

to recommend how compliance may be accomplished.

According to the National Water Quality Inventory 1986

Table 9-1. Membrane Concentrate Generation 

Membrane Process

Percent

Recovery of

Feedwater

Percent

Disposal as

Concentrate

UF 80–90 10–20

NF 80–95 5–20

Brackish water RO 50–85 15–30

Seawater RO 20–40 60–80

ED 80–90 10–20

146
Report to Congress (U.S. EPA, 1986a), 36 of 52 (69

percent) U.S. states and territories reported brine/salinity as a major ground-water contaminant, indicating that

proper disposal of concentrate should be a concern to

all professionals engaged in the use of membrane processes. Table 9-2 summarizes concerns and requirements

associated with conventional concentrate disposal

methods.

9.1.3 Early Disposal Regulations

The State of Florida has by far the highest percentage

of membrane process plants in the United States. A case

study of permitting requirement, disposal option, and

regulatory trends in Florida indicates the direction that

concentrate disposal regulation is likely to take in the

rest of the country. In Florida, regulatory agencies began

governing concentrate disposal in the late 1960s. The

relatively simplistic mass balance or dilution approach

taken in that era evolved to a more complex approach

by the mid-1980s. The trend towards more stringent

regulations has continued into the 1990s. This trend is

likely to occur in other states, such as California, Texas,

and states along the east coast where use of membrane

process plants is increasing.

Movement towards increasingly stringent regulation of

concentrate disposal has been brought on by regulations passed to preserve and protect surface water bodies, bird sanctuaries, endangered flora and fauna,

ground-water resources, and other environmentally sensitive areas. In particular, the permitting process has

become more difficult. In the late 1960s, no permit was

required to discharge concentrate to surface waters.

Later, the State of Florida required a permit, but did not

have a separate permit category for concentrate. Thus,

the State of Florida permitted its first concentrate discharge in the early 1970s as an industrial waste. In

retrospect, placing concentrate disposal under a separate waste permitting category would have been better

because EPA later issued a new permitting requirement,

the National Pollution Discharge Elimination System

(NPDES), that required discharges to surface waters to

be permitted. This resulted in dual permitting of concentrate discharges to surface waters.

As more recent membrane process technology began

being applied to potable water treatment, systems began being sited farther away from coastal surface bodies

of water and new disposal methods such as deep well

injection began being employed. The first spray irrigation of diluted concentrate occurred in 1978, and the first

deep well injection of concentrate occurred in 1983.

Between 1978 and 1983, regulatory requirements had

changed so much that only monitoring requirements

were imposed on the spray irrigation project, while very

stringent well construction integrity and monitoring requirements were required for the deep well injection

project.

9.1.4 Current Regulations and Their Trends

Regulations vary widely from state to state and, in some

cases, from county to county. In general, regulations are

designed to protect surface- and ground-water resources; rarely do they specifically address the disposal

of concentrates or brine. One specific regulation concerning the disposal of concentrate to ground waters

was passed in Florida on July 28, 1988. Passage of an

amendment to Chapter 17-28.700 of the Florida Administrative Code (FAC) allowed nonhazardous concentrates to be discharged through land application to

aquifers containing greater than 1,500 milligrams per

liter (mg/L) of total dissolved solids (TDS). More recently, increased public awareness, the activities of environmental organizations, and public concern have

increased the difficulty of siting and permitting membrane process plants in Florida. In recent years, bills

have been introduced in the Florida legislature to ban

Table 9-2. Concerns and Requirements Associated With

Conventional Disposal Methods

Disposal

Method

Regulatory

Concerns Other Requirements

Disposal to

surface water

Receiving stream

limitations

Radionuclides

Odors (hydrogen

sulfide)

Low dissolved

oxygen levels

Sulfide toxicity

Low pH

Mixing zone

Possible pretreatment

Multiple port diffusers

Modeling of receiving

stream

Deep well

injection

Confining layer

Upconing to USDWs

Injection well integrity

Corrosivity

Well liner

Monitoring well

Periodic integrity test

Water quality of

concentrate must be

compatible with the

water quality in the

injection zone

Spray irrigation Ground-water

protection

Monitoring wells

Possible pretreatment

Backup disposal method

Need for irrigation water

Availability of blend

waters

Drainfield or

borehole

Ground-water

protection

Monitoring wells

Proper soil conditions

and/or rock permeability

Sanitary

sewer

collection

systems

Effect on local

wastewater

treatment plant

performance (toxicity

to biomass or

inhibited settleability

in clarifiers)

None

147
deep well injection. To date, these bills have failed, but

one might be successful in the future.

After his election in 1994, Governor Bush of Texas

banned the use of deep well injection. At the same time,

Safe Drinking Water Act (SDWA) regulations are causing membrane processes to be declared best available

technology (BAT) for the removal of many contaminants.

Yet, without a viable method of concentrate disposal,

membrane processes cannot be used. If pretreatment

requirements drive the cost of disposal too high, some

proposed membrane projects will not be cost-effective.

Already, some cities have been unable to obtain disposal permits. At a recent meeting of the Florida Department of Environmental Protection (FDEP) Fact Finding

Committee on Concentrate Disposal, it was recommended that a permit application be made in advance

when alternative concentrate disposal methods will be

proposed; application before beginning the detailed design of a proposed membrane process plant is intended

to reduce losses if a plant cannot be permitted. It was

further recommended that the permit application be

made for the full plant nominal capacity required at

buildout. Given that regulations are a moving target, it

should be kept in mind that what is permittable now

might not be permittable in the future.

9.2 Conventional Disposal Methods

9.2.1 Surface Water Discharge

Disposal of brine is a critical issue facing WTPs throughout the country. Because the majority of surface water

discharges occur in Florida, the discussion that follows

is based on Florida’s experience.

The question of the environmental impact of brine disposal to surface waters was first addressed by the State

of Florida in 1986. At that time, the state determined that

concentrate disposal posed a significant threat to surface water quality and began requiring that any discharge to surface waters be permitted.

9.2.1.1 Constituents in Water Treatment Brines

Several constituents of WTP brines have been identified

as potential threats to surface water quality. In Florida,

ground water is the principal source of raw water.

Ground waters are devoid of dissolved oxygen and

contain hydrogen sulfide. This is a concern in relation to

surface waters, because a lack of dissolved oxygen and

the presence of detectable quantities of hydrogen sulfide are toxic to aquatic organisms. Fortunately, both

conditions are easily remedied via a relatively simple

and inexpensive treatment: aeration, which simultaneously adds dissolved oxygen and removes hydrogen

sulfide. Some facilities have also removed hydrogen

sulfide through chemical treatment. Typical Florida “endof-pipe” permits require a minimum dissolved oxygen

concentration of 5.0 mg/L and a maximum hydrogen

sulfide concentration of 0.04 mg/L.

The presence of radionuclides in the brine can also pose

a major water quality problem. Florida water quality

criteria contain limits for combined radium (Ra-226 +

Ra-228) and gross alpha. RO brines can contain combined radium at levels greater than 100 picoCuries per

liter (pCi/L)—much higher than the Florida surface water

standard of 5.0 pCi/L. Gross alpha levels are more

difficult to measure due to interference from the dissolved solids in brines. High gross alpha values, however, typically occur with high levels of combined radium

and should be handled similarly. The presence of radioactivity in any waste, including brine, can become an

issue of great public concern, increasing the need

to deal with radionuclide exceedances promptly and appropriately.

With brine, an additional disposal concern with surface

water discharge is the impact of the concentrated salts

(dissolved solids) on the receiving water. Water quality

criteria for fresh water surface waters greatly limit the

capacity of a fresh water body to receive brine. Unless

the water body is a stream or river with a flow many

times greater than the brine flow volume, discharge is

not likely to meet the surface water criteria, even after

mixing. Therefore, outfalls should be located to discharge brine waste to brackish or marine waters with

adequate flushing.

Fluoride and several metals have been found in high

concentrations in the brines of some facilities. Differences in the concentrations of these substances in

brines from different facilities reflect differences in the

composition of the ground water that provides the raw

water to the facilities. In some cases, fluoride and metal

concentrations in ground water exceed surface water

criteria; in other cases, they do not exceed these criteria,

but they are higher than the fluoride and metal concentrations in the receiving waters. When brine concentrations exceed receiving water concentrations, antidegradation issues must be addressed. The new metals

criteria proposed by EPA and adopted in Florida are

hardness-dependent for many metals. The ultimate impact of these changes has not been determined, but will

require careful examination.

Nutrients are a new area of concern in Florida. As

facilities have increased in size, their brine volumes

have also increased. Today, nitrogen and phosphorus

concentrations in large volume brines can provide a

significant nutrient load to receiving waters. Because

many Florida surface waters are already overenriched,

significant nutrient additions are not likely to be approved.

A permit to discharge brine to a surface water will require

establishing an effluent limitation—the brine quality nec148
essary to ensure that the receiving water quality is not

degraded beyond what is allowed by state and federal

regulations. While effluent limitations are typically developed through application of sophisticated mathematical

computer models, this is not always necessary for brine

discharges.

9.2.1.2 Dilution, Mixing Zones, Modeling, and

Multiport Diffusers

The hydrologic characteristics of receiving waters are

important in considering surface water discharges of

brines. The receiving waters’ flow is the first issue that

must be addressed. Most brine constituents are conservative materials that can be assimilated only through

dilution. If the discharge dominates the receiving water,

little dilution will occur. Small streams, for example, have

little ability to dilute a discharge. Neither do lakes (even

large lakes), because mixing and transport away from

outfalls are slow, resulting in a localized buildup of brine

constituents.

With the exception of aeration and hydrogen sulfide

removal, treatment of brine has not been practical in

Florida. Therefore, dilution has been the only way to

meet receiving water quality criteria that are not met at

the end of the pipe. Dilution calculations do not usually

require sophisticated models. Often, desktop mass balance calculations are sufficient to determine the necessary effluent limitations. This is especially true when the

surface water flow past the brine outfall is easily measured and an assumption of complete mixing is appropriate.

When an assumption of complete mixing is not appropriate, simple mixing models such as the EPA-supported

CORMIX1, CORMIX2, UPLUME, and UMERGE can be

used to establish effluent limits. The CORMIX models

are particularly attractive because they are compiled

from routines that the model chooses from based on

input specified by the user. EPA has identified methods

for determining tidal prisms, flushing times, and initial

dilution that can aid in the calculation of mixing areas

(U.S. EPA, 1985b). Many mixing models also permit

analysis of the use of a multiport diffuser, which is often

necessary to achieve adequate mixing in a reasonable

area. Models should be used that consider the number,

size, and orientation of multiple ports. Analysis is helpful

because multiple ports do not always provide better

mixing than a single port.

When concentrations in brine exceed surface water criteria, regulatory agencies may require the development

of a mixing zone. In Florida, for example, a mixing zone

is required for any discharge with end-of-pipe constituent concentrations in excess of numeric surface water

criteria. No implied opportunity for mixing is given. In

these cases, it is necessary to use a mixing model that

establishes geographic boundaries as well as numeric

constituent limits.

Assumptions specified in modeling exercises should be

conservative. That is, they should examine worst case

situations to ensure that the effluent limits developed will

protect receiving water quality and comply with permit

conditions. Minimum and maximum flows should be

modeled to examine the dilution provided and estimate

variations in brine plume boundaries under different conditions. These analyses require data on constituent concentrations in the receiving water and in the brine. If the

facility is not yet constructed, analyzing brine from a pilot

test facility might be possible. Otherwise, analysis of the

raw water and calculation of expected brine concentrations will be necessary.

9.2.1.3 Toxicity Relative to Surface Water

Discharges

Any discharge to surface waters must address the issue

of toxicity. Discharges are typically screened for toxicity

using whole effluent toxicity bioassay. These bioassays

expose sensitive test organisms to 100 percent effluent

for 48 (EPA) or 96 (some states) hours, while control

organisms are exposed to “clean” water. At the end of

the test, the number of deaths in the effluent is compared to that in the controls to determine if the effluent

is acutely toxic. If an effluent is found to be toxic in this

type of screening test, it is subjected to a definitive

bioassay in which the organisms are exposed to various

concentrations of effluent to determine the concentration lethal to 50 percent of the test organisms in the test

period.

For toxic effluents, the degree of toxicity determines

whether effluent limitations can be established, whether

a mixing zone will allow the establishment of effluent

limitations, or whether additional treatment to remove or

reduce the toxicity will be required. In some cases, the

facility may need to determine the cause of the toxicity

through a Toxic Identification Evaluation (TIE).

In Florida, bioassays are required for all permitted surface water discharges. The frequency with which these

tests must be performed varies with the volume of the

discharge and the type of discharge. Because several

discharges have failed bioassays, more test requirements are likely. The cause of brine toxicity has not been

definitively determined and might vary among facilities.

9.2.2 Disposal to Sanitary Sewers

Discharge of concentrates to sanitary sewer systems is

sometimes feasible if the concentrate mixture is not

toxic and does not adversely affect the clarifier settleability or restrict final effluent disposal. Generally, the

restrictions and requirements of sewer agency user ordinances must be followed. WTP agencies typically require detailed analytical screening before agreeing to

accept a proposed brine or concentrate discharge. Special attention will be given to the discharge’s effect on

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metal and hazardous constituent levels in WTP residuals, which must be disposed of safely.

In California and other western states, regional dedicated brine interceptor lines have been in place for many

years. Originally installed to carry away waste brines

from the oil industry and other industrial sources, they

are now being used for residual brines from municipal

well head treatment, such as brines from nitrate reduction ion exchange systems and concentrates from inland

desalters used to reduce TDS in ground waters contaminated by years of agricultural operations. The cost of

using these regional brine interceptors is often high—in

some cases, higher than the cost of the proposed treatment processes. Costs include contribution-in-aid to

construction of treatment facilities, other capacity

charges, construction of pipelines from the treatment

process to the interceptor lines, cost per acre or foot of

discharging to the interceptor sewer, and pretreatment

costs (if required).

9.2.3 Deep Well Injection

9.2.3.1 History of Deep Well Injection and

Injection Regulations

During the last two decades, intensifying awareness of

the problems associated with wastewater disposal into

the surface and near-surface portions of the earth (e.g.,

lakes, rivers, ponds, irrigation fields, and shallow aquifers) has increased interest in deep well injection. As its

name implies, deep well injection is the pumping of

waste into deep geologic formations. The injection depth

depends on site-specific geologic conditions and the

quality of the receiving waters as set by minimum criteria

in federal and local regulations.

Regulatory criteria for deep well injection of non-oil drilling waste waters was first formalized at the federal level

in a Federal Water Quality Administration order dated

October 15, 1970. That order only gave general guidance; it was not until 1973 that true criteria were established through the issuance of Administrative Decision

Statement No. 5 by the Administrator of EPA. After the

1980 amendments to the 1974 SDWA, EPA consolidated the rules into the underground injection control

(UIC) regulations and compiled them in the Code of

Federal Regulations (40 CFR Parts 144 and 146), with

general program requirements located in 40 CFR Part 122.

In May 1981, EPA developed a mechanism for granting

individual states primary enforcement responsibility, or

primacy, on UIC and issued a corresponding guidance

paper. Many states have since developed their own UIC

regulations, which, by law, can be stricter but not more

lenient than the federal rules. Most states have opted to

allow full regulatory control by EPA on certain types of

injection wells by choosing not to develop and enforce

rules that would give them primacy over those types of

wells. Regulations in 40 CFR Part 147 describe how

individual states operate under EPA rules.

9.2.3.2 Classes of Injection Wells

In the federal UIC regulations, injection wells are defined

as wells into which fluids are injected. A well itself is

defined as a bored, drilled, or driven shaft, or a dug hole,

with a depth greater than the largest surface dimension.

By definition, well injection means the subsurface emplacement of fluids through a bored, drilled, or driven

well, or through a dug well, where the depth of the dug

well is greater than the largest surface dimension (40

CFR Part 144.3). Specific inclusions covered by UIC

regulations are septic tanks and cesspools used for

hazardous waste disposal and those serving more than

20 persons per day.

Within this broad definition, EPA has classified injection

wells into five different groups as described in 40 CFR

Part 144.6 and 40 CFR Part 146.5. Paragraph 146.5

provides a detailed definition of Class V injection wells,

while paragraph 144.6 does not. Otherwise, the two

paragraphs are the same. The five classes of injection

wells are:

• Class I injection wells:

– Inject hazardous wastes beneath the lowermost

formation containing a source of drinking water at

least a quarter mile from the well.

– Inject nonhazardous industrial and/or municipal fluids beneath the lowermost formation containing a

source of drinking water at least a quarter mile

from the well.

• Class II injection wells:

– Inject nonhazardous fluids associated with production and storage of oil and natural gas.

• Class III injection wells:

– Inject fluids for the extraction of minerals by solution mining, leaching mining, etc.

• Class IV injection wells:

– Inject hazardous or radioactive water into or above

a formation that contains a source of drinking water

within a quarter mile of the well or that injects

hazardous or radioactive water into aquifers exempted from UIC rules by the EPA Administrator.

• Class V injection wells:

– All injection wells not included in the other four

classifications, including air conditioning return

wells, cesspools, cooling water return wells, drainage wells, recharge wells, salt intrusion barrier

wells, thermal recovery return wells, experimental

wells, etc.

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Under these five classifications, injection wells used for

the disposal of WTP residuals, such as concentrated

salt water from desalination, are classified as Class I

injection wells; within that classification, they belong to

Group 2, which includes nonhazardous industrial wells.

This industrial classification is significant because it invokes stricter requirements for well construction and

monitoring, even when the injection fluids are of better

quality than domestic injection fluids that can routinely

be disposed of into Group 1 Class I injection wells.

On the other hand, when WTPs are used to remove

hazardous waste from a source water, the residuals are

considered hazardous water and extra precautions must

be taken to ensure proper operation and monitoring of

the injection system. In some states, hazardous waste

injection wells are banned, and in several other states

with primacy rights, banning hazardous waste injection

wells is being considered at this time. In some cases,

chemicals used in the treatment process could be considered hazardous wastes. When these chemicals are

removed with the treatment residuals, they must be

neutralized or removed before they can be injected.

9.2.3.3 Water Treatment Plant Residuals

Two types of residuals are important: those derived from

the water being treated and those derived from the

chemicals being used to treat the water. Assuming a

noncontaminated source, residuals from water generally

consist of concentrated salt and minerals, the concentrations and types of which depend on the source water.

Added chemicals usually include antiscallants, acids,

flocculants, and pH adjusting chemicals; they may also

include nonhazardous cleaning solutions and other

chemicals.

Residuals can be suspended or dissolved. When deep

injection is considered, it is essential to evaluate the

effect of mixing these residuals with the receiving waters

in the injection zone. Mixing can cause the formation of

precipitates, flocculants, gases, and bacterial mats,

which can hinder the injection process and can harm the

receiving aquifer by plugging it or by otherwise reducing

its permeability.

9.2.3.4 Requirements of Injection Sites

An important element in the Class I Injection Well definition is the reference to injection below the lowermost

formation containing an underground source of drinking

water. An underground source of drinking water (USDW)

is any nonexempted aquifer or portion of an aquifer that

supplies water to a public water system (or is large

enough to supply a public water system) and that does

indeed currently supply drinking water to someone, as

long as that water contains less than 10,000 mg/L of

TDS. Thus, a Class I injection well must be sited such

that injection occurs into an aquifer zone with water

having at least 10,000 mg/L TDS (i.e., into waters that

are not USDWs). In addition, the injection zone of the

aquifer must be separated from the USDW zones above

it by hydrologically impermeable formations that preclude upward migration of the injected fluids into the

USDWs. These formations are referred to as confining

beds, although in reality they are never fully confining,

since, theoretically at least, every formation has some

measure of permeability.

A third requirement is that the injection zone be a good

receiving zone from both a water quality and hydrologic

point of view. From the water quality point of view, the

quality of the receiving water must not be aggressive to

the injection fluid. The two waters must be compatible

physically, chemically, and bacteriologically. From the

hydrologic point of view, the receiving zone must be of

high permeability and effective porosity such that the

volumes of water injected can be discharged into the

zone without excessively raising the pressure in the

receiving formation. Excessive injection pressure can

fracture the receiving zone and the confining layer

above it, causing upward migration of the injection fluids

into the USDWs. Similarly, fluid should not be injected

into areas where the confining zone has fissures or

fractures or is crossed by known faults, as these geologic features will become routes of injected fluid migration into the USDWs. In addition, injection should not

occur in areas where seismic forces are active; in some

cases, injection fluids have been suspected of acting as

a lubricant to induce slippage in underground formations, resulting in seismic disturbance at the land surface.

9.2.3.5 Environmental Concerns of Deep Well

Injection

The primary environmental concern with deep well injection is the potential for the injected fluids to migrate

into potable water zones. This migration can occur via

two mechanisms:

• Natural migration across confining layers that are not

as confining as they should be, in which case the

injection fluid migrates laterally into other potable

zones through geologic faults or by simply flowing

into other areas where the water is less than 10,000

mg/L TDS.

• Artificial migration of injected fluids through improperly constructed injection systems, such as defective

cement, construction induced fracturing near the

borehole, casing failures, etc.

Depending on the types of residuals being disposed, the

waste to be discharged will have a higher or lower

density than the receiving waters. Assuming a homogeneous aquifer, when the injection fluids are more dense,

they sink to the bottom of the injection zone, migrating

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first downward and then horizontally. When the injection

fluids are less dense than the receiving waters, they float

on top of the injection zone, forming a bubble of fresher

(less dense) fluids that produces an upward force across

the confining zone. Similar conditions, in reverse, occur

when the permeability of the injection zone is not homogeneous, (i.e., when the injection zone is more permeable in one direction than the other). For example, if the

permeability is higher in the vertical than in the horizontal direction, vertical buoyant forces will induce a vertical

movement of the injection fluid.

When vertical forces across the confining layer are induced by any of these injection pressures, vertical migration through the confining bed eventually occurs.

This may take years, hundreds of years, or longer, but

because no formation is totally impermeable, when the

data are entered into a formula, there will always be a

time when the fluid will move across the confining layer.

This environmental problem is often mentioned qualitatively, although quantitatively it may be of little or no

concern.

Two major concerns about confinement are uncertainty

about whether the confining layer extends over a sufficiently large area and uncertainty about whether the

confinement might be broken by faults, geologic caverns, or other geologic features. These issues are difficult to resolve because it is economically impossible to

drill enough test holes to show that the confining layer

exists everywhere over the area of concern. However, it

is logical to expect that a thick confining bed will not

suddenly thin out and disappear in a short distance.

Similarly, while it is possible to show the existence of

faults (typically identified through other sources in nearby

sites), it is not possible to affirm that no faults exist.

Environmental concerns related to well construction and

the problems associated with improper construction are

discussed in Section 9.2.3.7.

9.2.3.6 Regulatory Requirements

Because ground water is the primary source of water for

many communities, deep well injection is carefully regulated and monitored. Design and permitting are strictly

regulated by EPA when not regulated by the state under

primacy rules. Florida is a primacy state with a growing

practice of using deep injection wells for desalination

residuals disposal; it is a good example to describe how

the permitting process works.

Permitting Requirements

In Florida, permitting and regulatory control of deep well

injection is exercised by FDEP with aid from a technical

advisory committee (TAC). This committee consists of

five voting members who represent environmentally oriented government agencies. The voting members of the

TAC are representatives of the regional office of the U.S.

Geological Survey (USGS), EPA’s Atlanta office,

FDEP’s Tallahassee office, the appropriate local Water

Management District office, and the local FDEP district

office. Two nonvoting members representing the local

county pollution control office and the local county health

department are also part of the TAC. The representative

of the local FDEP district office serves as the TAC

chairperson.

The TAC reviews and approves every engineering design and exercises control of every aspect of construction and testing from beginning to end. The TAC also

reviews and approves specified construction materials

and methods, including the size and weight of the steel

casings, the cement and cement additives used in the

casing installation, the method of drilling, the frequency

and types of tests to be conducted, and other details.

The TAC meets regularly with injection well engineers

and owners to discuss project specifics, including review

and approval of the depth of installation of each casing

and the tests used to ensure mechanical integrity of the

well. A preliminary Plan of Study is usually submitted to

the TAC to inform the regulatory agencies about the

project and the proposed method of construction and

testing.

After extensive review, the TAC may suggest changes

that are incorporated into a final Plan of Study. A formal

permit application is then made for a test well which, if

successful, usually becomes the first injection well of the

system. In areas of Florida where the suitability of geologic conditions for deep well injection is in doubt, the

test well may be preceded by an exploratory well, which

is usually converted into a monitor well.

Under the Florida Administrative Code, there is a 90-day

permitting clock for the permit application procedure.

However, a permit is rarely issued within 90 days of

application submission because the permitting clock

stops each time a TAC member requests additional

information. When the TAC chairperson is satisfied with

the application, FDEP issues a notice of receipt of a

complete application. This notice is published in the

local newspaper and a public hearing date is set and

advertised.

If there is no public opposition to the application at the

hearing and no one requests additional public hearings,

the TAC will begin reviewing of application and all associated data, including design documents and specifications. At the next regularly scheduled meeting of the

TAC, voting members may give approval for the project

to proceed with the test well or they may again request

additional changes. After approval by the TAC, FDEP

issues a second notice of intent to permit, which is also

published in local newspapers, and another public hearing is advertised. If there is no opposition, FDEP will

issue the permit to construct and test the test well. This

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permit usually lasts a year. At the end of the operational

test period, the inspection well operator must apply for

a final operating permit.

Design Requirements

Specifications and contract documents for permit applications must follow the design criteria required by regulatory agencies. These criteria require the use of:

• Well head vacuum release or air release valves to

prevent fluid column separation and air entrapment.

• Hydropneumatic tanks to prevent water hammer.

• Cushioned valves and pump systems designed to

maintain injection velocities below 8 feet/second

(ft/sec) to prevent erosion in the well bore.

• Instrumentation and controls to maintain injection

pressures below the fracturing limits of the formation

receiving the treated effluent.

Designs must include a sufficient margin of error to

ensure that the formation’s ability to receive the effluent

is not exceeded and that the opportunity for aquifer

recovery is provided. The system must be designed to

minimize water quality changes of injected fluid to avoid

turbidity and air entrapment, to avoid formation of precipitates, and in general to ensure that nothing clogs the

formation.

Construction requirements may include use of:

• A drilling pad capable of retaining spilled fluids during

construction to avoid surface contamination.

• Multiple casings to protect each fresh water producing zone that is drilled.

• Casing and liner materials compatible with the injected fluids.

• Heavy wall casing (e.g., a minimum of 0.375-inch

casing wall thickness).

• Sulfate-resistant cement only, a minimum of 2.5

inches of grout between exterior casings, and a minimum of 5 inches of grout outside of the injection casing.

• Appropriate annular fluid between the injection casing

and the liner of the wells such that the packer in the

annular space is not subject to chemical attack by

that fluid.

• Backflow preventers and check valves.

• Well head designs that facilitate monitoring and logging without major disturbance to system operation

and without requiring a drill rig on site.

Finally, monitoring, inspection, and testing—both during

construction and during operation of the injection system—must be addressed in the design process. Principal elements include:

• Performing geophysical logging before and after each

casing installation.

• Conducting downhole television surveys before casing installation and after final casing is installed.

• Sampling water quality at every change of formation

or every 30 feet to detect the 10,000 mg/L TDS zone.

• Sampling and testing the quality of water in the injection zone, monitor zones, injection fluid, and shallow monitor wells.

• Monitoring annular pressure and the quality of annular fluid (industrial wells only).

• Conducting aquifer performance tests, injectivity

tests, and mechanical integrity tests.

• Measuring and continuously monitoring and recording injection pressure and injection volume.

• Conducting recertification testing every 5 years.

9.2.3.7 Well Construction

A well must be properly constructed so that it serves its

function without harming the environment. Proper construction ensures that a well passes construction-related

and site condition-related mechanical integrity requirements. Of these, only construction-related mechanical

integrity requirements can be controlled through proper

construction, use of adequate materials, and proper

operation and maintenance. Site conditions must be

controlled by proper well siting, which usually occurs

after the first well in a given area has been drilled to

show that the site is adequate. Even in areas where

other injection wells are operating satisfactorily, any new

well should be nominated as a test injection well because of the possibility that the construction could turn

out to be faulty or the site inadequate.

Contractor Selection

Selecting a qualified well driller for a project is best

accomplished by prequalifying drillers based on their

experience and reputation and by inviting qualified firms

to bid on the specifications and drawings. Special attention must be given to ensuring that drillers have the

proper equipment to complete the contract on schedule.

The lowest responsible bidder is generally awarded the

contract. Private firms may choose to negotiate contracts rather than select a driller through competitive

bidding. Bid price is influenced by a number of conditions, especially the volume of work available at the time

of bidding.

Drilling Pad and Drilling Equipment

The first stage in drilling the deep injection well involves

drilling the shallow portion of the hole and setting the

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surface casing. By specifying a small drilling rig during

this stage, costs can be minimized.

Once the surface casing is installed, an impervious

drilling pad must be constructed to prevent contamination of the aquifer by salt water spills during construction.

The pad is constructed of 4- to 6-inch thick reinforced

concrete and is surrounded by a berm. All drilling equipment, fluid circulation equipment, and settling tanks are

installed on top of this drilling pad. Shallow monitor wells

drilled around the pad serve as an early warning system

to detect salt water spills. As an additional precaution

against spillage, a temporary flow control device is installed before drilling through salt water zones.

After the surface casing is set and the pad and shallow

monitor wells are complete, a large drilling rig is brought

on site. The well size and depth will determine the type

of rig. Assembling a rig and installing all peripheral

equipment can take up to a month.

Fluid Circulation System and Other Equipment

During drilling, all drilling fluid is recirculated through a

series of settling tanks to allow collection of rock cuttings

from the hole. Provisions are sometimes made for temporary storage of salt water overflow in polyvinyl chloride (PVC) lined ponds. This salt water can later be

reinjected into the well. Cement and cement additives

for grouting casings usually are brought to the site dry

and are stored until they are mixed and pumped by a

cement company specializing in well grouting.

Field offices of the engineer and drilling contractor are

generally located on site. Once drilling starts, it usually

continues 24 hours a day, 7 days a week. Construction

supervision should be provided around the clock by the

consulting engineer.

Geologic Sampling—Proper Well Siting

Formation samples are collected at preselected intervals and at every change of geologic formation. Samples are catalogued according to their physical

characteristics and mineral composition to identify confining layers and injection zones.

Several formation cores must be obtained from specific

depths in the well. The best core samples typically come

from the confining zones, since samples from the injection zones are usually broken and fractured. After field

inspection and cataloguing, the cores are shipped to a

laboratory where they are tested for physical and chemical characteristics, including permeability to water.

These tests determine if the confining beds are sufficiently impermeable to meet regulatory standards.

Construction Procedures

All drilling is preceded by a pilot hole 8 inches or less in

diameter, which yields better formation samples and

facilitates more accurate geophysical logs. The pilot

hole is reamed to full size after it has reached total depth

for casing installation. Full size is usually about 6 inches

larger than the nominal diameter of the casing to be

installed. The one exception is the reamed hole for the

innermost (injection) casing. This hole is always 10

inches larger in diameter than the injection casing to

maintain the minimum requirement of 5 inches of cement grout in the annular space around the injection

casing

Casing Installation

Injection well construction requires several sizes of casings, with the largest diameter being used at the surface

and successively smaller casings being used deeper

into the ground. The smallest is the injection casing at

the bottom of the well.

A drilling rig elevator and an auxiliary elevator are used

to raise each casing to the top of the drilling platform and

then to lower it into the hole. Special care should be

exercised to prevent damage during installation. As sections of casings are lowered into the well bore, they are

welded into place. To ensure that there are no voids in

the welds, two full welding passes should be performed.

Inspectors should examine each weld and pressure

tests should be conducted to detect any imperfection. If

a defect is found, corrective action can be taken before

the well is put into operation. The welding operation

usually continues around the clock over a period of

several days. The pressure test on the final string of

casing must be conducted at 1.5 times the estimated

injection pressure when the well is completed and

placed into operation.

Cementing Operations

After the casing has been installed, cement must be

placed in the annular space between the walls of the

well hole (or inside a telescoping casing) and in the

casing itself. A sulfate-resistant cement (Type II, Class

H) is generally used because deep aquifers contain

unusually high levels of sulfate. Cementing is accomplished by pumping cement through a pumper truck into

a small diameter “tremie” pipe in the annular space at

the bottom of the well hole. As the annular space fills

with cement, the tremie is slowly pulled out of the hole.

The first cement installation (lift) usually contains no

additives. Subsequent cement lifts usually contain various concentrations of additives to give the cement more

or less plasticity, faster or slower setting times, or

greater or lesser density to fill voids or cavities in the

formation. The cement mix and additives are selected

by the cement engineer and the consulting engineer

after a review of geologic samples and geophysical logs.

The local regulatory program is responsible for and must

review and approve the cement program. If no local

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program exists, EPA must review and approve the cement program prior to implementation.

Water Sampling

Throughout the drilling process, representative water

samples must be collected from discrete sections of a

formation to define water quality at specific depths in the

hole. Samples are collected using a packer consisting

of a perforated pipe with two inflatable rubber seals at

each end. When the seals are inflated and the packer

assembly is pumped, a water sample is obtained from

the formation section between the seals. These samples

are pretreated at the site and shipped to a laboratory for

analysis. Time-dependent parameters are measured in

the field.

Injection Test

The injection test is performed by injecting clean water

into the injection well and monitoring pressure changes

and water quality. Desalination reject water cannot be

used for injection tests because system failure could

result in ground-water or USDW contamination.

Performing an injection test requires that the monitor

well be completed and equipped with gauges to monitor

pressure changes—typically, a fluid-filled gauge backed

by a mercury manometer that measures pressure in

inches of mercury or 10-inch and 12-inch test grade

gauges. Electronic gauges are usually avoided because

they can be unreliable in injection test conditions. Use

of pressure transducers installed at the bottom of the

hole is ideal because they eliminate the need to correct

for pipe friction losses and water density differences. All

pressure readings are corrected to feet of water and are

referred to a fixed datum, usually the surface of the

drilling pad. This calculation facilitates monitoring of injection pressure and well injectivity over time.

Samples of the water being injected are collected at

various intervals during the test to determine the quality,

density, and temperature of the water.

After the well is constructed and the injection test is

performed, a downhole television survey of the well

should be conducted using a slim television camera

equipped with a fish-eye lens and a light source. The

videotape of the downhole survey is used to confirm that

there are no holes in the casing and no defective joints.

The videotape also provides a permanent visual record

of the injection zone before injection occurred so that a

comparative analysis of the casing and injection zone

can be performed in the future.

Other geophysical surveys are usually conducted at the

same time to complement those conducted during construction.

Well Head Completion and Monitoring

Proper design of the well head will facilitate collection of

all future monitoring data. Easily readable gauges at the

injection and deep monitor wells should always be

specified and, if funding is sufficient, remote reading

transducers are highly recommended.

The most important design elements are the pressure

and vacuum release valves because they are the ultimate line of protection against high pressure and vacuums that can develop during electrical malfunctions or

scheduled and unscheduled shutdowns. A hydropneumatic tank system should also be added to all wells to

further increase protection.

Proper valve design also ensures safe access to the well

when geophysical logs and television surveys are conducted in the future. For example, proper valve design

permits insertion of probes (including the television camera) and avoids spills of injected residuals.

9.2.3.8 Operational Considerations

Any injection system operating program must reflect the

hydrogeological characteristics of the injection zone and

the volume and chemistry of the injection fluids. The

injection rate determines the injection pressure, which

may not exceed the maximum value for the injection

system. The maximum allowable injection pressure is

the highest pressure that:

• Preserves the integrity of the formations in the injection zone and the overlying confining zone.

• Prevents significant change in the fluid movement

capabilities of the overlying confining zone.

• Protects the mechanical integrity of the well structure.

Similarly, the permissible injection rate for a deep injection well is that which:

• Protects the mechanical integrity of the well structure.

• Preserves the integrity of the formations in the injection zone.

• Preserves the fluid movement capabilities of the injection zone.

Injection Pressure and Preservation of Formation

Integrity

Safe operating pressures can be determined with certainty only by conducting fracture tests on the formations

in the injection zone. The maximum safe bottom-hole

injection pressure typically ranges from about 0.5 to 1.0

pounds per square inch (psi) per foot of well depth,

depending on the types of materials that make up the

formations and geological conditions in the open hole.

In Florida, for example, a value of 0.6 psi per foot of well

depth is typically used to calculate the maximum safe

155
bottom-hole injection pressure for a 2,700-foot deep

injection well into a dolomitic formation. Multiplying 0.6

per foot of well depth by the well depth of 2,700 feet

yields a maximum safe bottom-hole injection pressure

of 1,620 psi.

The maximum safe well head injection pressure is the

difference between the maximum safe bottom-hole injection pressure and the measured bottom-hole hydrostatic pressure. If the system in the example above

produced a bottom-hole hydrostatic pressure of 1,200

psi during pumping at a design capacity, the difference

between the maximum safe bottom-hole injection pressure and the measured bottom-hole hydrostatic pressure would be 420 psi (1,600 to 1,200 psi). Dividing 420

psi by a safety factor of 3.0 yields a maximum well head

injection pressure of 140 psi for the well. The safety

factor might differ from site to site because it is an

engineering designer’s choice, not a regulatory requirement. Friction losses are then calculated for the design

flow and the diameter of the injection liner. High friction

losses are subtracted from the 140 psi to yield the

maximum safe well head injection pressure to be used

in monitoring reports.

Injection Pressure and Prevention of Fluid

Movement in the Confining Zone

Success in preventing significant change in the fluid

movement capabilities of the confining zone is somewhat difficult to evaluate due to different interpretations

of the word “significant.” Analyses of cores and geophysical logs from the confining zone should show low

permeability values. Regardless of how small, however,

any permeability value entered into a leakage equation

will result in some leakage value. Determining what is

or is not significant leakage is subjective.

Injection into a well generates piezometric heads in all

radial directions. The amount of leakage should be calculated within the area determined to be critical—usually the quarter-mile radius or area of review (AOR) from

the injection well. Beyond the 1/4-mile radius of the

AOR, the differential pressure across the confining zone

should also be calculated to determine when pressure

buildup will (at least in theory) produce leakage beyond

the AOR.

Injection Pressure and Protection of Well

Mechanical Integrity

The maximum well head injection pressure for a deep

injection well must also protect the mechanical integrity

of the well structure. In particular, it must be lower than

the bursting pressure of the steel pipe used for the inner

(injection) casing and the bursting pressure of the injection liner within the inner casing. The bursting pressure

of water filled, unsupported pipes of various diameters

and wall thicknesses is provided by the manufacturer.

Because the steel pipe is always encased in cement

grout the entire length of the well, the maximum permissible pressure inside the casing, is much higher than the

bursting pressure of the unsupported casing. As a result,

it is not usually a significant consideration in design,

construction, or operation. For the liner, the collapsible

pressure rather than the bursting pressure is more important, especially during construction. For example, an

unsupported fiberglass casing with a bursting pressure

of 600 psi might withstand only 100 psi in compression.

Thus, the use of a 600-psi injection pressure must be

weighed against the need to keep the annular pressure

below 100 psi.

The annular packer also influences mechanical integrity.

Packers should be tested to hold in tension and in

compression.

Injection Pressure Data Reporting

A data reporting program must be part of operation and

maintenance (O&M) procedures. The program should

include regular reporting of operational well head injection pressures. If at any time the injection pressures

show a rapid increase toward the maximum, measures

should be taken to alleviate an impending violation before it occurs.

If, under normal operating conditions, the well head

injection pressure suddenly starts to approach the permitted maximum, the injection pumps and deep injection

well should be shut down until an emergency evaluation

can be conducted and other contingency plans put into

operation. On the other hand, if the well head injection

pressure rises result from excessive injection rates, the

injection should be throttled back immediately to a rate

that will reduce the well head injection pressure to below

the maximum.

Injection Rate and Protection of Well Structure

Due to the potential for damage, maximum injection

rates cannot be established through empirical injection

rate tests. Instead, standard practice has been to establish a maximum permissible injection velocity of 8 ft/sec,

the maximum velocity that can be expected to not create

erosion of the open hole in most formations. This rate

has been adopted as the ruling criterion for maximum

safe injection rate in some states.

The 8-ft/sec velocity limit protects the injection well from

large flow velocities that can lower fluid pressure close

to the saturated vapor pressure at the injection pump (or

at any other flow obstruction in the pipeline, such as a

sudden turn or constriction). Lowering fluid pressure

within the injection line can lead to cavitation within the

pump and the injection line, resulting in fluid flow interruptions that can damage the system. Because higher

injection rates within a well require larger injection velo156
cities, maintaining lower rates minimizes the possibility

of damaging pressure reductions.

Injection Rate and Preservation of Formation

Integrity

Stable operating conditions also help protect the integrity of the injection zone. Pressure fluctuations under

variable injection rates create variable stress-strain relationships within formations, particularly the rock

shelves that separate the porous zones. These stresses

and strains in turn can induce rapid cyclical flexing of the

rock matrix, which can affect the rock matrix and fracture

the formations. To prevent fracturing, rapid cyclical injection should be kept to a minimum. This can be accomplished by using the hydropneumatic tank and pressure

and vacuum release valves discussed above.

Injection Rate and Protection of Injection Zone

Capabilities

The transmissivity of the injection zone and its ability to

receive injection fluid must also be considered. The

maximum rate of injection that can be received by the

injection zone is largely limitless as long as the well

structure and the rock matrix within the injection zone

are able to receive the fluid. Nevertheless, economic

considerations favor the use of the lowest possible injection pressure, since the frictional head loss increases

exponentially with increasing flow rate.

In general, therefore, the limiting criterion for increasing

injection rate is a matter of economics and the need to

protect the well structure and injection zone. The maximum safe injection rate for a deep injection well is

determined by the most restrictive of those considerations, usually the 8-ft/sec limitation on fluid velocity

within the injection pipe.

9.2.3.9 Monitoring Requirements

During the time that a deep injection well is operational,

some types of monitoring must be conducted on a regular basis and others must be carried out on an asneeded basis.

Regular Monitoring

Regular Injection Well Monitoring. Regular monitoring of

a deep injection well consists of:

• Keeping continuous record of the variable nature of

the flow rate for each 24-hour period and recording

the daily flow volume. Data on the daily flow rate and

the total monthly flow, together with the average and

maximum daily flow and the maximum instantaneous

rate, must be reported monthly.

• Recording the variable nature of the well head injection pressures. Daily average and daily maximum

injection pressures must be recorded for each 24-

hour period and reported monthly.

• Recording the annular space pressure to show any

fluctuations with flow rate. The annular space volume

contracts and expands with injection pressure

changes as well as with the temperature changes of

the injection fluid.

• Running and reporting a specific injectivity test at

pre-approved intervals to ensure that the receiving

formations are not being affected by excessive plugging or excessive dissolution.

• Monitoring the water quality of the WTP residuals in

accordance with required parameters in the regulations. Water quality must be reported monthly, but is

usually reported by the plant rather than the well

operator.

A sample report format (adapted from forms used by

FDEP) is shown in Figures 9-1 through 9-5.

The injected fluid must also be tested weekly as part of

the operation of the well itself. These tests involve taking

a sufficient quantity of injection fluid from the pipeline

between the WTP and the deep injection well. For analysis for TDS, total suspended solids (TSS), chloride,

sulfate, conductivity, pH, and temperature. The temperature reading and the pH must be taken immediately after

the sample has been collected. The remaining water

quality parameters may be determined later in the water

quality testing laboratory. An ongoing record of the various parameters must be maintained.

All of the data must be kept on site and available for

inspection by regulatory authorities on request. The log

sheets must include any annotations and will constitute

the record of operation.

At the time of sample collection or sample analysis, the

following monitoring information must be provided:

• Date, exact place, and time of sampling or measurements.

• Person responsible for performing the sampling or

measurements.

• Date(s) analyses were performed.

• Person responsible for performing the analyses.

• Analytical techniques or methods used.

• Results of the analyses.

Regular Monitor Well Monitoring. Monitor wells are usually used to monitor two water-producing zones above

the injection zone: one at the first permeable zone above

the injection zone and one at the first USDW above the

confining zone. As part of overall ground-water monitoring, those zones must be tested on a monthly basis and

the results reported to regulatory agencies.

157
Figure 9-1. Monthly operational report, page 1.

158
Figure 9-2. Monthly operational report, page 2.

159
Figure 9-3. Monthly operational report, page 3.

160
Figure 9-4. Monthly operational report, page 4.

161
Figure 9-5. Monthly operational report, page 5.

162
Monthly monitoring of the monitor zones must include

the following steps:

• Recording the pressure or water level in each zone,

noting monthly daily highs and lows on the monthly

form.

• Collecting a monthly water quality sample from each

of the two monitor zones (every 2 weeks for the

shakedown test period).

The water quality samples collected monthly (or bimonthly at first) must be taken to a water quality testing

laboratory to measure TDS, chloride, sulfate, and specific conductances (conductivity). Temperature and pH

should be measured in the field immediately after sample collection. Records of monitoring information should

include:

• Date, exact place, and time of sampling or measurements.

• Person responsible for performing the sampling or

measurements.

• Date(s) analyses were performed.

• Person responsible for performing the analyses.

• Analytical techniques or methods used.

• Results of the analyses.

The results from the water quality analyses must be

submitted on a monthly basis. A typical report format is

shown in Figures 9-1 through 9-5.

Special As-Needed Monitoring

Special injection well monitoring may consist of the

following tests, either alone or in combination: mechanical integrity test, pressure test, television survey, noise

and radioactive tracer test, and controlled injection test

(injectivity test).

The mechanical integrity of a deep injection well must

be checked every 5 years via a 1-hour pressure test on

the injection casing. To perform this test, a temporary

packer must first be installed at the bottom of the injection liner. Within 3 months, the results of the pressure

test must be submitted to regulatory authorities with an

interpretative letter. Copies of the results must also be

retained at the plant office.

A television survey of the deep injection well must also

be made every 5 years. This involves running a television camera down to the maximum depth possible

within the open hole. Within 3 months of the survey, a

copy of the videotape must be submitted with an interpretative letter, and a copy of the videotape must be

retained on site. Similarly, a radioactive tracer test

and/or a noise log may also be required.

If regular monitoring records begin to indicate significant

increases in well head injection pressures over time, a

controlled injection test and/or television survey may be

required before 5 years to ascertain the probable cause.

If a sudden significant increase occurs in well head

injection pressures, regulatory agencies must be notified within 24 hours. A testing program could then be

required for the deep injection well, which would probably include a controlled injection test and a television

survey. After the results from these tests are analyzed,

an interpretative report must be submitted within 3

months. If the tests results require any work in the deep

injection well, the results of the work must also be

submitted within 3 months.

A routine injectivity test must be performed every month

during the shakedown period. Subsequently, a special

injectivity test must be performed quarterly or as

needed. For meaningful comparisons between tests,

standard test procedures must be followed exactly each

time the test is conducted. These procedures include:

• Once a month, allowing the final effluent storage tank

to fill sufficiently to allow the injection system to operate at the maximum permissible injection rate.

• Measuring the shut-in pressure in the well (prior to

the start of the injectivity test).

• Connecting the injection pump directly to the injection

well (bypassing the hydropneumatic tank) and operating it at the predetermined number of revolutions

per minute that will result in the maximum permissible

injection rate.

• After 1 hour of injection at the maximum permissible

rate, recording the injection pressure at the well head.

This must be repeated again just prior to the end of

the test, at the end of the second hour.

• Arranging the two pressure readings together to determine the specific injection pressure for the injection rate used.

• Reporting the specific injectivity index as the ratio of

the fixed volumetric rate to the specific pressure (the

average of the two readings). This result is the specific injectivity index (in gallons per minute/specific

pressure).

• Preparing a running chart by plotting the locations of

the monthly results. In the chart, the x and y coordinates represent the volumetric rates and their pressures, respectively.

Sampling Procedures

Water quality monitoring of desalination plant residuals

is discussed in Section 10.3.3.1. Sampling procedures

for the deep monitor well require samples to be collected

163
after the system has been properly evacuated. The

methodology must be customized for each well.

Shutdown and Startup Procedures

Normal Operation. Once put into operation, injection

wells operate automatically without the need to open or

close any valves. The air release and vacuum release

valves on the well head are the only valves that open

and close with any regularity, and they do so automatically. Moreover, the hydropneumatic tank acts as a

buffer and creates redundancy so that even the automatically operated air and vacuum release valves need

not be controlled in any way. In fact, the control valves

to the air release and vacuum release valves should

always be in the open position whether or not the well

is in operation. The valve that connects the release

valves to the well should always be open as well.

Startup. Before the well is put into operation and when

it is shut down, all of the valves should be closed. Once

the injection pump is turned on, the master valve (a

slow-closing butterfly valve that connects the well to the

plant) should be opened within 30 seconds to prevent

buildup of pressure in case the hydropneumatic tank is

bypassed. If the hydropneumatic tank is not bypassed,

the capacity available in the tank reduces the need to

open the master valve immediately. Even then, however,

the master valve must be opened before the hydropneumatic tank fills (within 1 to 2 minutes).

Shutdown. Shutting down the well is a straightforward

operation, requiring only that the master valve be closed.

If the shutdown is temporary, other valves may be left

open. If the shutdown is long-term, all valves except the

vacuum and pressure valves should also be closed.

Attaching a lockable chain to all valves to prevent tampering is strongly recommended. Alternatively, the valve

handles can be removed. The master valve should also

be specially protected by a chain lock or by removing its

handwheel. Valves used for water sampling should be

shut down at all times except while sampling. The handles should be removed when not in use to prevent unauthorized opening, which would waste the annular fluid.

Schedules and Procedures for Calibration and

Operation of Pumps and Monitoring Equipment

Calibration procedures depend primarily on the types of

equipment used and the manufacturers’ recommendations. Each equipment manufacturer has specific procedures that are documented in reference manuals; these

should be followed. Single copies of these manuals are

provided by the manufacturer through the contractor

and are to be retained on site at all times.

9.2.3.10 Data Reporting

Data reporting requirements for a deep injection well

include:

• Regular monthly operational reports

– Pressures and volumes (injection well)

– Water quality (injection well)

– Water quality (upper and lower monitor zones)

– Pressures (upper and lower monitor zones)

• Special interpretative reports

• Noncompliance reports

Regular Monthly Operational Reports

Operational reports for each calendar month should be

submitted no later than the 15th day of the following

month or according to permit requirements.

Pressures and volumes (Injection well). This part of the

monthly report (see Figures 9-1 and 9-2) follows the

EPA-suggested reporting format for injection wells.

Page 1 of the monthly operational report contains four

sections. Sections I and II are self-explanatory. Section

III details restrictions on the maximum daily injection

volume, maximum injection rate, and maximum well

head injection pressure. The maximum and average

daily injection volumes and the cumulative total volume

for the month are recorded in Section III-A. Sections III-B

and III-C are used to record the maximum, average, and

minimum injection rates and the well head injection

pressures for the month. Section IV contains instructions for filling out the report.

The maximum daily injection volume during the operating period is the largest of the average daily injection

volumes calculated for each of the operating periods

during the month. The present average daily volume for

the month is the total volume injected for the month

divided by the sum of the lengths of the operating periods during that month. The total volume injected to date

is taken directly at the end of the month from the daily

volumes recorded.

The maximum injection rate during the month is the

largest of the daily maximum injection rates. The minimum operational injection rate is the lowest of the daily

minima. The average injection rate during the month is

equivalent to the present average daily volume recorded

in Section III-A-3, except for a time that includes only the

length of actual injection.

The maximum well head injection pressure during the

month is the largest of the daily maxima. The minimum

operational well head injection pressure is the lowest of

the daily minima. The estimated average well head injection pressure during the month must be estimated for

each operating period from the daily values. The time164
weighted average is calculated from the various estimated average well head injection pressures for the

month.

Column 1 on page 2 is used to record the start, end, and

length of each operating period during the month. The

maximum and minimum injection rates must be read

each day and recorded in Column 2. The average injection rate is calculated from the daily total flow at the end

of each day, dividing the difference by the duration of

operation that day. Both the maximum and minimum

well head injection pressures are recorded each day in

Column 3. The average daily well head injection pressure is also estimated each day and recorded in Column

3. The total volumes are recorded at the end of each day

whenever operating periods exceed 1 day. If the operational period is less than a day, the flow at the end of the

period is recorded.

Water quality (Injection well). Page 3 of the monthly

report (see Figure 9-3) is used to list the laboratory

results of tests performed on samples of the injection

fluid after it has passed through the desalination treatment plant (just prior to injection). The report form provides space for recording the results of the tests

recommended for inclusion in the monitoring program.

The water quality samples should be sent to a certified

water quality testing laboratory, which will return the test

results to the operator for inclusion in the monthly operational report.

Water quality (Upper and lower monitor zones). Page 4

of the monthly operational report (see Figure 9-4) includes data from both the upper and lower monitor

zones. It covers the range of parameters described

earlier in this chapter. Data analysis must be performed

by a certified laboratory.

Pressures (Upper and lower monitor zones). Page 5 of

the monthly operational report (see Figure 9-5) should

be filled in whenever water level measurements are

taken. Water level data for both monitor zones should

be recorded daily. The water level of the lower monitor

zone can be measured by taping down to the water

surface and recording the distance (as depth to water,

in feet) below the measuring point (MP). Alternatively, a

calibrated recording bubbler system installed on the well

can be used as long as the calibration holds true. The

water level for the upper monitor zone can be read

directly from the gauge or recorder.

Interpretative Reports

Interpretative reports are usually submitted following

any type of special monitoring. These reports typically

consist of letters; the specific format depends on the

nature of the special monitoring. Interpretive reports are

usually developed by the engineering consultants for the

WTP and are submitted within 3 months of completion

of the special monitoring.

Noncompliance Reports

Noncompliance reports are required in the case of abnormal events (e.g., equipment breakdown, power failure, clogging of the well, etc.). In such cases, a verbal

report must be made within 24 hours, a preliminary

written report must be submitted within 3 days, and a

comprehensive written report must be submitted within

2 weeks. These reports must describe the nature and

cause of the event, the period of noncompliance (dates

and times), steps taken to correct the problem and to

prevent its recurrence, emergency procedures in use

pending correction of the problem, and the estimated

date by which the facility will again be functioning in

accordance with permit conditions.

If the deep injection well is abandoned, the owner and/or

operator must provide adequate documentation indicating that the well was properly abandoned. The documentation must describe the method of abandonment

and evaluate the results of the operation.

9.2.3.11 Stimulation Program

An injection zone intersected by the open hole of a

properly operating deep injection well derives its ability

to receive injection fluids from natural cavities in the rock

matrix and from larger interstices. These cavities are

relatively large and numerous and are not usually susceptible to clogging, unlike the smaller cavities within the

rock matrix itself. Therefore, an injection zone is not

considered to be susceptible to any serious clogging

effects, although it is possible that injectivity will decrease over time. Observing the restrictions that apply

to the quality and composition of the injection fluid will

reduce the chance of clogging and maintain the high

injectivity essential to successful operation of the well.

Federal regulations require that a stimulation program

be available in case it is needed. If the deep injection

well becomes clogged or plugged and thus becomes

less receptive to the injection fluid, it may be necessary

to enact a well stimulation program. Unless the cause

of the reduced intake is already known, the well operator

might also be asked to conduct a special monitoring

program. An interpretative report on the deep injection

well must then be submitted, discussing the problem

and reporting on the special monitoring program. The

report should recommend corrective actions (including

a stimulation program for the deep injection well) to

rehabilitate the well.

A well stimulation program may include surging, acidizing, or a combination of both. Other stimulation techniques (shooting, hydraulic featuring, and vibratory

explosions) are used primarily within the petroleum in165
dustry, are controversial, and are only recommended as

a last resort.

Surging to stimulate the deep injection well involves use

of a pump that can inject water into or discharge water

from the well at continuously fluctuating rates. This action is designed to loosen any matter that might be

clogging the rock openings. Well acidization is designed

to chemically break down any matter clogging the rock

pores, thereby improving flow movement through the

injection zone. The acid used in this procedure should

be 18°Baume (27.92 percent) hydrochloric (muriatic)

acid. The acid should contain a nontoxic stabilizer to

prevent after-precipitation of dissolved minerals and a

nontoxic inhibitor to prevent corrosion of well head components. Use of the stabilizing and inhibiting chemical

agents must first be discussed with the appropriate

regulatory agencies.

The procedure used for acid treatment is subject to

regulatory approval. The two most common methods are:

• Method 1: Full strength acid is introduced to the well

by means of a 3/4- or 1-inch diameter drop pipe. The

acid must not be introduced by pouring it from the

top of the well casing. The drop pipe must be black

iron or plastic pipe, not galvanized pipe. Overflows

and vents required for safety must be provided. The

acid is introduced to the well starting from the bottom

of the open hole. The drop pipe is gradually retracted

during filling operations to ensure that equal concentrations of acid are present in the entire length of

open hole. The quantity of acid used should be 1-1/2

to 2 times the volume of water in the open hole

section of the well. At no time during the acidizing

process should the pH of the well contents be allowed

to rise above 2.0. Additional acid is introduced to the

well as required to ensure that the pH remains below

2.0 for the duration of treatment. After introducing the

acid to the well, the acid should begin being agitated

as soon as possible. Agitation can be accomplished

by surge block bailers or by any other appropriate

means available. The acid should be agitated for 3

to 4 hours. Following acid treatment, the well should

be pumped clear of all sediment until the water is free

of discoloration and the pH level returns to normal

(between 6.0 and 8.0). Pumped out waste should be

discharged to the plant.

• Method 2: If pressure acidizing is used instead, pressure control is provided by a sealing arrangement

between the well casing and the acid drop pipe. The

seal must be capable of maintaining a minimum operating pressure of 100 psig in the well. The sealing

arrangement should include the pressure gauges and

valves needed to accurately determine the gauge

pressure in the well and to ensure that acid and air

can be introduced into or vented from the well. Pressure should be maintained at 75 to 100 psig for a

minimum of 2 hours. Following pressure acidization,

the waste valve should be opened and the well

pumped at a rate to produce substantial drawdown

to clear the well of all sediment until the water is free

of discoloration and the pH level returns to normal

(between 6.0 and 8.0). Discharge should be to the

plant. Following the initial acid treatment, a second

acid treatment might be required to further increase

the well yield.

The second method is generally used only if the clogging is so extensive that high pressures can be maintained in the injection zone during acid application.

The results of a stimulation program must be tested by

a special monitoring program. Both the stimulation program and the special monitoring program should then

be fully described and analyzed in an interpretative

report submitted to regulatory agencies.

9.2.3.12 Costs for Concentrate Disposal

The estimated cost of deep well disposal (cost per gallon

of membrane treatment plant capacity) is shown in Table

9-3. The actual cost of well construction is shown in

Figure 9-6. Well construction costs depend primarily on

well depth and diameter. Nevertheless, the injection

liner material and the manner of liner installation can

cause the cost of a typical well to vary by as much as

$50,000.

In addition to the O&M costs shown in Table 9-3, a

recurring cost is associated with the mechanical integrity

test that must be conducted every 5 years. Contractor

costs for this test usually range between $70,000 and

$90,000; engineering fees add another $50,000 to

$60,000. Permit renewal fees can also be high. In Florida, the fee for renewal of a Class I well permit is

$10,000.

9.2.4 Boreholes

Drainfields and boreholes can be used with regulatory

approval when the concentrate water quality meets the

regulatory requirements for discharge into surficial

ground-water aquifers and when soil conditions and

Table 9-3. Concentrate Disposal Costs

Deep Well

Disposal

Facility

(mgd)

Capital

Cost

($/gallon)

Engineering

Cost ($ per

gallon)

Land

Requirements

(acre)

O&M

Cost

($/1,000

gallons)

 3 $0.58 $0.087 0.5 $0.032

 5 $0.44 $0.066 0.5 $0.024

10 $0.40 $0.060 1.0 $0.022

15 $0.37 $0.056 2.0 $0.02 

20 $0.30 $0.045 3.0 $0.016

166
permeability of rock strata permit. In Florida, this occurs

in coastal areas where ground-water quality exceeds

10,000 mg/L TDS and USDW regulatory parameters are

not exceeded. Surficial aquifer discharge of brine from

coastal seawater RO plants that use shallow boreholes

or seawells (where the ground water is very brackish)

can be more cost-effective and environmentally palatable than an ocean outfall.

9.2.5 Spray Irrigation

In general, spray irrigation is viable only if the following

conditions are met:

• The TDS, chloride content, or salinity of the concentrate or the mixture of concentrate and a blend liquid

do not exceed a level that will damage the grass or

crop being irrigated.

• There is a requirement for irrigation water in the vicinity of the WTP to avoid a long conveyance system.

• A backup disposal system (e.g., storage for use during sustained periods of rainfall, when irrigation is not

needed) is available.

• The water quality of the surficial ground water is protected from degradation.

• Local, state, and federal regulations are met.

• A system of monitoring wells is in place to check

overall irrigation system performance. A monitoring

plan is generally a prerequisite to obtaining an operating permit.

If these conditions can be satisfied, spray irrigation can

be an attractive option because it is economical, provides an appropriate source of irrigation water, and conserves natural resources.

Major variables affecting the cost of spray irrigation

include land, concentrate conveyance systems to the

disposal site, and site developments (e.g., relocation,

land preparation, surface runoff control, subsurface

drainage, distribution and irrigation, storage requirements, and pretreatment).

Spray irrigation of concentrates almost always requires

dilution prior to irrigation to prevent pollution or degradation of ground-water resources while meeting regulatory requirements. Dilution also reduces the chloride

Figure 9-6. Well drilling cost.

167
content to acceptable levels to prevent damage to the

grass or crops being irrigated.

Disposal of concentrates by spray irrigation is being

proposed for several very large systems currently being

designed. Recently, a sprayfield demonstration program

in Boynton Beach, Florida, was funded jointly by the

South Florida Water Management District and the City

of Boynton Beach. The primary purpose of the demonstration program was to better manage water resources

by using a product ordinarily disposed of as waste.

Years after the permitting for the Boynton Beach concentrate disposal was finalized, deep well injection was

chosen as the disposal method.

9.3 Nonconventional Methods of

Concentrate Disposal

9.3.1 Evaporation and Crystallization of

Brines for Zero Discharge

9.3.1.1 Brine Characteristics

Achieving a zero liquid discharge at any desalination

facility almost always requires some type of evaporation

and crystallization process. Evaporation is differentiated

from crystallization only in that crystallization usually

involves the development and processing of slurries,

while evaporation refers mostly to liquid solutions. Techniques for handling high-density slurries are quite different from those for liquids, however, and are always more

expensive per gallon of water evaporated. Figure 9-7

shows a basic zero discharge flow sheet.

Brines from desalination operations generally become

wastewater because they cannot be concentrated further in membrane systems due to a high TDS content

or the presence of scale-forming materials. Calcium

sulfate and silica are the usual scaling components,

although the barium and strontium sulfates can also

contribute to scaling. Fortunately, techniques for dealing

with scale-forming materials have been developed for

use with both evaporators and crystallizers.

Control of scaling is generally accomplished through

“seeding.” Both evaporators and crystallizers use the

same fundamental approach, but the concentrations of

solid particles differ greatly. In either case, sufficient

crystals, or “seed slurry,” are maintained in suspension

to provide enough crystalline surface area to effectively

allow most or all of the precipitation to occur on existing

crystals rather than on the heat exchange surfaces. This

is a simplistic explanation, but it is correct in concept and

will help guide the process designer in developing workable zero discharge systems.

Just as TDS limit the operation of membrane systems,

they can limit the operation of evaporators and brine

concentrators. Crystallizers can process streams with

virtually any TDS level, even up to the point at which the

circulating stream is more a melt than a solution. As an

example, crystallizers are often used to concentrate

sodium hydroxide to 50 percent by weight or higher

(dissolved) while crystallizing sodium chloride from the

liquor.

9.3.1.2 The Brine Concentrator

The workhorse of the zero discharge plant, the brine

concentrator is most often designed as a vertical tube

or falling film evaporator, although horizontal spray film

and plate type evaporators are also used in this application. Various manufacturers provide brine concentrators; currently, most are used in the electric utility

industry. These units concentrate cooling tower blowdown and other wastes to about 20 percent total solids,

which are then processed further to a solid waste in

either solar ponds or a crystallizer. The first brine concentrators entered commercial service at coal-fired utilities in the mid-1970s, and some of these units are still

in service today.

The common characteristic of most brine concentrator

designs is the circulation of a slurry of calcium sulfate

crystals, which act as seeds. Calcium sulfate and other

scale-forming compounds preferentially precipitate on

the circulating seed crystals (rather than the heat transfer surfaces), preventing scaling. The development of

the seeding technique for calcium sulfate and silica

allowed brine concentrators to process wastewaters at

or near saturation in calcium sulfate and silica to very

high concentrations. Up to 20 percent total solids by

weight is often achievable in the brine concentrator

discharge, while a distillate of more than 10 parts per

million (ppm) TDS can be returned to the power station

for use as boiler makeup. Figure 9-8 shows the typical

components of a brine concentrator. Because most

Figure 9-7. Zero discharge. brine concentrators are powered using a mechanical

168
vapor recompression (MVR) cycle, the system shown in

Figure 9-8 includes a compressor. Direct steam power

could be used, but the cost of steam energy almost

always exceeds that of MVR, except in special situations.

MVR evaporators or, more simply, vapor compression

(VC), use a compressor to raise the condensing temperature and pressure of the evaporated water sufficiently to reuse the vapor as the primary heat source for

evaporation. In this manner, the heat of vaporization in

the evaporated water is recovered immediately in the

evaporation of more process water. In a typical steam

driven system, the heat of vaporization in the evaporated water is simply lost to the final condenser. At best,

only a portion of the heat is recovered in a multiple-effect

evaporation system.

In MVR evaporators, the compressor draws suction

from the evaporating water in the vapor body of the

evaporator. The condensing temperature and pressure

are raised in the compressor 10° to 15°F above the

boiling temperature in the evaporator. The now higher

temperature vapor is then fed to the outside of the heat

transfer tubes, where it gives up heat and condenses.

More water evaporates inside the tubes, which is in turn

recompressed, and so on. The steam is reused without

first condensing it.

MVR systems can achieve very high thermal efficiencies

compared to steam driven systems. MVR brine concentrators require evaporation energy that, if converted to

steam, is equivalent to 1 pound of steam evaporating

over 28 pounds of water; for example, only 35 British

thermal units (BTUs) are required to evaporate 1 pound

of water.1

 A single-effect evaporator, in contrast, requires

1 pound of steam per pound (or less) of water. In terms

of energy consumption, MVR is at least 28 times more

efficient than a single-effect steam driven evaporator.

Representative brine concentrator process conditions

for zero discharge applications are listed in Table 9-4.

These conditions are shown as ranges and should be

used as guidelines. Many factors affect the operating

limits of brine concentrators, but a few are important for

planners of zero discharge systems.

Crystallizers are much more expensive to build and

operate than evaporators or brine concentrators. Because of this great cost differential, operating a membrane system at its limits of concentration might become

counterproductive in some instances. If the membrane

system waste brine is too highly concentrated, a brine

concentrator might be unusable (see Table 9-5), and a

much larger crystallizer will be required to reach zero

discharge. The upper limit for feed concentration into a

brine concentrator is determined by the need to keep

enough seed material in suspension to prevent scaling

as well as by the solubility of the more soluble salts. This

usually occurs with a feed TDS of somewhere between

Figure 9-8. Brine concentrator system (HPD, 1994).

1 During operation, 85 kilowatt-hours (kWh)/1,000 gallons (gal) evaporated is routinely obtained; 85 × kWh 3,418 BTU/kWh/(1,000 gal ×

8.3 lb/gal) = 35 BTU/lb.

Table 9-5. Effect of Concentration Factor (CF) on Calcium

Sulfate Seed Concentrationsa

Ion/Species Feed 5 CF 10 CF 20 CF 50 CF

Calcium 400 2,000 4,000 8,000 20,000

Sodium 1,100 5,500 11,000 22,000 55,000

Chloride 550 2,750 5,550 11,000 27,500

Sulfate 2,500 12,500 25,000 50,000 125,000

TDS 4,550 17,510 33,460 65,360 161,060

Total solids 4,550 22,750 45,500 91,000 227,500

Suspended solids

(calcium sulfate)

 5,240 12,040 25,640 66,440

a All values in ppm.

Table 9-4. Typical Brine Concentrator Process Conditions in

Zero Discharge Applications

Feed TDS 2,000–20,000 ppm

Feed temperature 40–120°F

Preheater approach temperature 7–15°F

Concentration factor 8–120

Seed slurry concentration 1–10%

Boiling point rise 1–10°F

Compression ratio 1.20–2.0

Recirculation pump rate 20–40x feed

TDS of waste brine, weight % 15–22%

Total solids waste brine 15–30%

Overall power/1,000 ga l75–110 kWh

Distillate TDS 5–25 ppm

169
25,000 and 40,000 ppm, a range easily reached with

membranes.

Boiling point elevation (BPE), which increases with concentration, also limits the economical concentration factor. The BPE is the difference in boiling temperature at

constant pressure between pure water and a solution.

In an MVR system, the compressor must first overcome

the BPE before any evaporative energy can be recovered. Typical commercial systems operate with BPEs

ranging from 1.0° to 4.0°F. Above about 5° to 7°F, energy

requirements begin to rise significantly.

Seeding

A brine concentrator cannot concentrate into the precipitation region of soluble salts such as sodium sulfate,

sodium chloride, or the mixed salt glauberite,

CaNa2(SO4)2. This upper limit of concentration is typically reached with brackish waters at the 20- to 25-percent TDS level. Because most brine concentrators must

be seeded to prevent scaling, however, sufficient seed

material must be maintained in suspension in the system. The seed material most often used is calcium

sulfate because it is the primary scaling component of

most brackish wastewaters and is, therefore, usually

present.

If insufficient seed is maintained in suspension, scaling

of the heat transfer surfaces promptly results due to the

deposition of calcium sulfate, silica, or both. Scale removal is expensive and time-consuming, and generally

cannot be tolerated on a regular basis. Operating experience has shown that calcium sulfate seed crystals can

effectively seed for themselves, for other sulfates such

as barium and strontium, and for silica. In most applications, the seed material is formed in situ from dissolved

calcium sulfate present in the feedwater.2

To precipitate most of the calcium sulfate and maintain

a useful seed concentration in the brine concentrator,

the feedwater must be concentrated many times. As a

practical minimum, the operating brine concentrator

must maintain a seed suspended solids concentration

of at least 1 to 1.5 percent by weight calcium sulfate or

10,000 to 15,000 ppm suspended solids. With a feed

completely saturated in calcium sulfate, the minimum

concentration factor that will result in enough seed material, even with a seed recovery system, is about eight

concentrations, or 87.5 percent water recovery. Fifteen

or more concentrations are preferred, but operation at

eight concentration factors is possible if a seed recovery

system is used.

The seed recovery system recoups seed material in the

discharge from the brine concentrator and recycles the

naturally produced seed material. The apparatus generally consists of a hydrocyclone bank or settling tank that

recovers suspended solids from the waste brine blowdown and recycles them back into the system. For

example, consider a feedwater with the characteristics

shown in Table 9-5, where calcium sulfate is close to

saturation. The data in Table 9-5 show the effect of

concentrating the feedwater to various degrees on suspended solids and TDS. Even with a feedwater near

saturation with calcium sulfate, 20 concentration factors

(20 CF) are needed to reach 2.5 percent or 25,000 ppm

seed concentration in the evaporator if no seed recycle

is used.

This example illustrates why it is necessary to achieve

the maximum cycles of concentration in the brine concentrator. Because the brine concentrator is practically

limited to a maximum TDS of about 200,000 ppm and

needs at least ten concentration factors to reach moderate seed concentrations, the upper feed TDS limit for

a brine concentrator is in the range of 25,000 to 40,000

ppm TDS. For example, concentrating a feedwater of

25,000 ppm for eight cycles will result in total solids of

200,000 ppm, near the maximum. With a 35,000-ppm

feedwater, eight cycles results in 280,000 ppm, beyond

the limit of most brine concentrators. Yet eight cycles are

required to develop sufficient seed material to prevent

scaling without an external supply of seed. Thus, a

feedwater with 35,000 ppm TDS may be a candidate for

a crystallizer rather than a brine concentrator.

When silica is the primary scaling component, or when

calcium sulfate is not near saturation, addition of exogenous calcium sulfate might be required. This can be

accomplished by adding calcium chloride and sodium

sulfate or sulfuric acid and lime. Conceivably, anhydrous

calcium sulfate could be added directly to the evaporator, although this has not been done commercially.

Another reason to design brine concentrators at the

maximum concentration factor obtainable is that sensible heat loss in the concentrated waste discharge

stream becomes a consideration as recovery decreases. Waste is hot and filled with slurry, making

conventional heat exchangers difficult to operate. The

additional cost required to recover the heat in the waste

is usually not worth the cost of providing a larger compressor and accepting the energy loss penalty. At a high

discharge rate (e.g., a waste flow equal to 20 percent of

the feed flow), however, the compressor can have difficulty making up the heat loss, and makeup steam might

be required.

In summary, brine concentrator designs with greater

than 87.5 percent recovery (8 concentration factors) can

usually be made to function in the normal operating

range. Below 87.5 percent recovery, special considerations are required to maintain both proper seeding and

a balanced heat input.

2 At startup, of course, calcium sulfate is added directly to the brine

concentrator to form the initial charge of seed material.

170
Design Considerations: Major Components

As shown in Figure 9-9, the main components of a brine

concentrator are:

• Distillate heat feed heat exchanger: Usually a titanium plate type heat exchanger, this recovers heat

from the hot distillate into the feedwater. Typical approach temperature for the feedwater is 10°F below

the boiling point.

• Deaerator: Feedwater deaeration is required to prevent calcium carbonate scaling, to eliminate noncondensible gases that would interfere with heat transfer,

and to remove dissolved oxygen for corrosion control.

Ordinarily, the deaerator is a packed column (316L

shell and internals, ceramic or 316L packing) with

steam supplied from a hot distillate flash tank or a

compressor bleed stream.

• Main heater: Wastewater is inside the tubes, and

steam is outside. A falling film vertical tube unit is

shown. Typically, 2-inch outside diameter tubes are

used. Vertical lengths range from 30 to 55 ft. Tube

material is usually titanium grade 2, with a 0.028- to

0.035-inch wall. Noncondensible venting occurs at

approximately 1 to 3 percent of steam flow. Design

heat transfer coefficients range from 400 to 600

BTU/hr/ft2

/°F for vertical tubes. Horizontal tube units

are usually operated with the wastewater sprayed on

the outside of the tubes. Typical tube diameter is 1 inch.

• Vapor body: The main vessel holding the circulating

waste water also provides disengagement of steam

from the liquid. The vapor body can be integral with

the heater or a separate vessel. Typical volumes

based on feed retention time range from 30 to 60

minutes of feed flow. The vapor disengagement area

usually sets the minimum diameter of the vapor body.

Materials used depend on the equilibrium concentrations of chloride and magnesium; 316L and 317L

have been used with success in relatively low chloride applications with very good deaeration. Inconel

625 and SMO 254 or equivalent alloys have seen

more use since the late 1980s.

• Demisters/Entrainment separators: Mesh pads or

chevron-style entrainment separators are used to

clean the vapor stream of entrained droplets of very

salty water. Correct design is very important in MVR

applications to prevent corrosion and buildup in the

compressor. Entrainment separators can be external

to the vapor body and used with vapor scrubbers if

very high purity distillate is required. Chevrons and

mesh pads are usually fabricated in the same material as the vapor body.

• Recirculation pump: Circulation of the concentrated

waste stream through the tubes is required. Typical

circulation rates in brine concentrators are 20 to 40

times the evaporation rate.

• Compressors: With low BPEs (e.g., below 4° to 5°F),

single-stage centrifugal compressors are widely

used. Typical compression ratios for single-stage

units are 1.2 to 1.4. For higher BPEs, lobe-type compressors or multiple-stage centrifugals are needed.

Compressors as large as 5,000 horsepower have

been used in MVR installations. Guidevanes or variable speed motors are required to provide efficient

operation over a wide range of inlet flow rates. ComFigure 9-9. Brine concentrator capital and operating costs.

171
pressor wheels are fabricated in stainless steel or

cast iron. Housings are typically cast iron. Good entrainment separation is critical to prevent corrosion

and/or wheel buildup.

• pH control: pH adjustment is usually needed to convert carbonate and bicarbonate in the feedwater to

carbon dioxide before removal in the deaerator.

• Scale control: Scale control compounds are injected

just before the feedwater preheater. This delays the

precipitation of calcium sulfate at the higher temperatures found in the preheater and the deaerator.

• Startup boiler: In an MVR system, it is necessary to

raise the circulating water temperature to near the

boiling point before engaging the compressor. The

required startup steam is furnished by the plant or by

a small electrically fired startup boiler. For example,

with a 200-gallon per minute system, a 200 kw

startup boiler could bring the system to operating

temperature in 10 to 15 hours.

Design Considerations: Process

For the desalting engineer, the most important process

consideration is a system design that delivers feedwater

within the known operating constraints of the brine concentrator system. This typically requires:

• A calcium sulfate concentration greater than 50 percent of saturation in feedwater. This assures that

enough will be available to produce seed crystals.

With less than 50 percent of saturation (about 750

ppm calcium sulfate) and less than 15 concentration

factors, external seed supplies may be necessary.

• A TDS level low enough to permit high concentration

factors in the brine concentrator without precipitation

of soluble salts. Eight to 10 concentration factors are

the minimum required for normal operation.

Volatile components of the feedwater will distill over in

the brine concentrator and contaminate the recovered

water. Typical volatile components encountered include:

• Ammonia: Carryover is a function of pH, but even at

pH 6 to 7, ammonia can be found in the distillate.

• Boron/Boric acid: Boron carries over less than ammonia and is less sensitive to pH conditions.

• Organic acids: These are highly sensitive to pH and

species. Higher pH results in less carryover as the

ionized forms dominate.

• Trihalomethanes, etc.: The degree of carryover must

be determined by laboratory or pilot testing.

• Hydrogen sulfide: Carryover is a function of pH. Oxidation to sulfate might be necessary.

Capital and Operating Costs

The graphs shown in Figures 9-9 and 9-10 provide a

rough estimate of the capital and operating costs of

typical brine concentrator systems. For smaller systems

below 100-gallons per minute capacity, the equipment

is generally skidded. Larger systems require some field

fabrication of the vapor bodies and assembly of the

structural steel. The costs shown in the graphs are

based on installed outdoor systems with foundations.

Control rooms are assumed to be located within the

central plant facilities.

9.3.1.3 Waste Crystallizers

Crystallization technology for zero discharge grew out of

the chemical process industries. It was extended to the

nuclear power industry with the development of mixed

salt waste crystallizers for ion exchange regeneration

wastes. There it was necessary to produce a nearly dry

solid from the low-level wastes. The forced circulation

crystallizer has now become the primary tool for production of dry or nearly dry waste salts for zero discharge.

A basic waste crystallizer flow scheme is shown in Figure 9-11. In this illustration, an MVR cycle is shown, but

steam can be used instead. Figure 9-12 shows a simplified steam system with a barometric condenser. Because the crystallizer circulates a slurry, some method

of dewatering is required. The drawings show a centrifuge, but a filter press, rotary filter, or belt filter could also

be used.

Crystallizers do not have the same feed supply limitations as brine concentrators; streams with widely varying composition can be effectively processed. Because

the crystallizer does not generate a liquid blowdown,

concentration factor is not the appropriate term to describe system operation. The crystallizer simply evaporates water, leaving a circulating crystal slurry. The slurry

is dewatered at a rate that maintains a slurry concentration in the crystallizer of about 25 percent suspended

solids by weight. The TDS level eventually reaches an

equilibrium point, determined largely by the exact composition of the feed stream.

Residual water is entrained in the dewatered crystals,

but the amount is small compared to the inlet flow.

Centrifuges produce solids with 10 to 20 percent entrained liquid, while filters typically leave 40 to 50 percent entrained liquid with the solids. The entrained liquid

can be as high as 50 percent TDS; thus, in many applications, the hot crystals can absorb all the residual water

into water of hydration, leaving a dry-to-the-touch solid

at room temperature. For example, sodium sulfate exists as the anhydrous form at typical atmospheric boiling

temperatures, yet forms the decahydrate (Na2SO4-

10H2O) at room temperature, absorbing more than its

weight in water.

172
Design Considerations: Process

Typical operating conditions for a forced-circulation

waste crystallizer are shown in Table 9-6. With regard to

the feed stream, all of the dissolved and suspended

material will remain in the solids. The overall system

concentration factor for the desalting process, brine concentrator, and crystallizer can be very high. In some

applications, it may be possible to concentrate relatively

low levels of metals or other components in the feed to

levels that create a concern. As in a brine concentrator,

volatile materials will transfer to the distillate.

Steam is generally required if the waste solution has a

high percentage of magnesium chloride or nitrates.

These compounds are highly soluble and exhibit high

BPEs. At equilibrium, the solution boiling point might

well exceed the compression ratio available in typical

MVR systems. Steam at 25 to 30 psig or higher will be

needed. A condensing water supply equal to about 40

to 50 times the crystallizer feed rate will also be required.

In many cases, it should be possible to use the brine

concentrator and the membrane system product water

to condense the crystallizer vapors in a direct contact or

barometric condenser.

By necessity, the MVR waste crystallizer will require a

high lift compressor. A multistage or positive displacement compressor will also be required. Above a compression ratio of about 2.0 to 3.0, steam begins to be

favored from an economic viewpoint in almost all cases.

Figure 9-10. Brine concentrator cost components.

Figure 9-11. Waste crystallizer system (HPD, 1994).

Table 9-6. Typical Brine Crystallizer Process Conditions

Feed TDS >20,000 ppm

Feed temperature 40–220°F

Concentration factor N/A

Boiling point rise 1–50°F

Compression ratio (MVR) 2.0–3.0

Recirculation pumping rate >200x feed

Solids, % free liquid hot 15–50%

Solids, % free liquid cool 0–10%

kWh/1,000 gal (MVR) 150–200 kWh

BTU/1,000 gal (steam) 10 mm BTU

kWh/1,000 gal (steam) 2–4 kWh

Cooling water (steam) 50x feed

Distillate TDS 15–50 ppm

173
Because the waste crystallizer usually has a relatively

small capacity, steam might be preferred for simplicity of

control and maintenance, even in cases with low BPEs.

For large capacity systems (100 gallons per minute or

greater), the designer or specifier can consider the use

of multiple-effect systems to lower steam usage. Multiple-effect systems use less steam than single-effect

units, but have higher capital costs.

Separating solids from the crystallizer circulating liquor

can be difficult. Centrifuges are almost always effective,

but they are expensive to buy and maintain. In order of

solids dewatering effectiveness and cost, options for

solids removal include:

• Centrifuge (highest cost, lowest residual water).

• Belt filter (automatic operation, 40 to 50 percent retained liquid).

• Filter press (manual or semiautomatic, 40 to 60 percent retained liquid).

• Hydrocyclone (lowest cost, no moving parts, 50 to 70

percent retained liquid).

If solids meeting the “no free water rule” are required,

testing of the filter options and hydrocyclone might be

needed to demonstrate sufficient absorption of the free

water.

Design Considerations: Major Components

Vapor release and entrainment separation determine

the minimum size of the vapor body. A minimum of 6 feet

of straight side above the liquid level is required for

proper disengagement. Maximum vapor release velocities can be estimated using the following equation:

Velocity (ft/sec) = k[(Dl

 – Dv)/Dv]

0.5

where

Dl = density of the liquid in lb/ft3

Dv = density of the vapor in lb/ft3

k = constant ranging from 0.10 to 0.04

(dimensionless)

The minimum diameter can easily be calculated. For

example, a 50-gallon per minute crystallizer operating at

atmospheric pressure with a 20-degree BPE will require

a minimum diameter of 10.3 ft with k set at 0.05. Vapor

body materials must be suitable for harsh service. Inconel 625 is used for rad waste units, while titanium,

Inconel 825, and SMO254 or FRP units have been used

in commercial service.

The recirculation pump must overcome the heater pressure loss and the lift to the surface in the vapor body.

The vapor body must be located enough above the heat

exchanger to suppress boiling in the heat exchanger

and to provide sufficient NPSH. Pump heads, in net

TDH, typically run at about 15 to 25 feet. The circulating

liquor specific gravity is greater than 1.2. Axial flow

pumps are usually specified, but low (lower than 1,100)

revolutions-per-minute centrifugal pumps can also be

used.

Cooling water or air cooling can be used in the condenser (which is not required for MVR). The cooling

requirement, in BTUs/hr, is approximately equal to the

steam BTUs/hr. Direct contact condensation using the

Cooling 


Water 


Inlet

Barometric


Condenser

Distillate
 Solids Recovery Hot Well

Return

Main Heater


(Two Pass)

Steam


Inlet

Entry


Nozzle Vapor


Body

Entrainment


Separator

Feed


Inlet

Centrifuge

Recirc.


Pump

Steam Driven Forced Circulation Crystallizer

Vapor Space

Figure 9-12. Steam driven circulation crystallizer.

174
desalted product can be considered as well. Tubes are

usually 304 or 316 ss.

Tubes in shell units are typically used for the heat exchanger. Its orientation can be vertical or horizontal.

Boiling inside the tubes should be suppressed with sufficient head above the exchanger and tube velocities

must be sufficient to avoid settling of the solids; 6 to 8

ft/sec is typical. Temperature rise per pass through the

heat exchanger is kept low to avoid violent flashing in

the vapor body. Titanium tubes are standard. Differential

temperatures (steam to liquor) of 20°F can be used.

Heat transfer coefficients are a function of temperature

and concentration. For preliminary sizing at atmospheric

boiling, an overall coefficient of 400 BTU/°F/ft2

 can be

used.

Capital and Operating Costs

Figures 9-13 to 9-16 provide ways of estimating the

capital and operating costs of crystallizers. The costs

are based on single unit systems at 100 percent of

design capacity. Single-effect steam driven systems are

assumed. The values provided range roughly an order

of magnitude for installed systems.

9.3.2 Evaporation Ponds

The suitability of evaporation ponds for concentrate disposal depends on ideal climatological conditions. The

principal factors that affect the evaporation rate are

relative humidity, wind velocity, barometric pressure, air

and water temperature, and the salinity of the water.

The need for dry, arid conditions where evaporation

losses are much greater than the amount of rainfall limits

use of evaporation ponds to just a few areas in the

United States. In addition, the amount of land required

Figure 9-14. Steam power crystallizer cost components.

Figure 9-15. MVR crystallizer capital and operating costs.

Figure 9-13. Steam power crystallizer capital and operating costs.

175
is an economic consideration. Because evaporation

losses are directly proportional to area, large areas of

land must be available for this option to be viable. A

backup system is necessary, since an unusual weather

event could affect the evaporation rate. Some large

brackish water RO plants in the Middle East have employed this method of concentrate disposal successfully

or have used it in combination with percolation onto the

desert.

9.3.3 Emerging Technologies

Currently, no identifiable new technologies are predicted

to be used in disposing of brines. Refinements, increased

efficiencies, and combinations of existing disposal technologies might lower the cost of disposal, but no new

options are anticipated.

9.4 Costs Associated With Brine Waste

Disposal

Capital and O&M costs for concentrate disposal will be

specific to a given site. Cost variables include:

• Quality of concentrate water.

• Quantity of concentrate water.

• Disposal method.

• Distance from disposal point (and the length of the

conveyance system).

• Pretreatment required, if any.

• Regulatory requirements for method of disposal.

• Permitting.

• Use of energy recovery.

• Requirement for backup or redundant disposal method.

• Climatological conditions.

• Residual pressure of the concentrate stream exiting

the membrane process plant.

9.5 Conclusion

Concentrates from membrane process plants have

been disposed of for over two decades without incident.

Membrane processes are important treatment technologies that will assist the water treatment industry in meeting present and future drinking water regulations.

Membrane technologies probably will be the best available technology in terms of the most organic and inorganic contaminant removal for the amount of capital

invested. Safe methods of concentrate disposal as well

as fair and pragmatic regulations will be necessary for

membrane processes to continue. It is the responsibility

of all those involved in the membrane process industry

to work together to:

• Prevent damage to the environment that could occur

from improper disposal of concentrate wastes.

• Ensure that proper engineering judgment is used in

interpreting agency rules and regulations.

• Guarantee that future regulations do not unduly limit

the use of membrane process technology.

• Secure new methods for concentrate disposal through

research and development.

Figure 9-16. MVR crystallizer cost components.

176
Chapter 10

Radioactive Waste Disposal

Nearly all of the radionuclides found in drinking water

supplies are naturally occurring and are members of one

of three natural radioactive decay series: the uranium

series, the thorium series, and the actinium series. In

1976, the U.S. Environmental Protection Agency (EPA)

promulgated interim drinking water regulations for radionuclides that established limits for gross alpha activity, gross beta and photon activity, and the specific

elements of radium-226 (Ra-226) and radium-228 (Ra228) (40 CFR Part 141).

On July 18, 1991, the Agency proposed new regulations

that would increase radium maximum contaminant levels (MCLs) and add limits for uranium and radon (40

CFR Parts 141 and 142). The final regulation was

scheduled to be promulgated April 30, 1995; as of this

printing, however, this promulgation date is postponed

until an undetermined time after December 1995. EPA

is considering separating the other radionuclides from

radon because regulation of radon is under a congressional freeze.

10.1 Background

A variety of treatment processes exist to remove radioactive contaminants from drinking water, such as conventional coagulation/filtration, ion exchange, lime

softening, reverse osmosis, and granular activated carbon (GAC) adsorption (U.S. EPA, 1976b; Brinck et al.,

1978; U.S. EPA, 1977; Lassovszky and Hathaway, 1983;

Sorg et al., 1980; and Lowry and Brandow, 1985). These

processes separate the contaminants from drinking water

and concentrate them in the waste streams. Because

these processes are also commonly used to remove

other nonradioactive contaminants, they could potentially concentrate significant levels of radioactivity in

waste treatment residuals even if the treatment was not

originally designed or intended to remove radioactivity.

In addition to the waste streams containing the concentrated contaminants, some materials used to treat

drinking water adsorb radioactive contaminants and permanently retain them. One good example is sand used

in filtration processes. Sand adsorbs and retains radium

on its surfaces. Although the radium does not interfere

with the treatment process, at the end of its useful life

the sand must be replaced and disposed of.

The handling and disposal of the water treatment plant

(WTP) residuals containing naturally occurring radionuclides pose significant concerns to water suppliers, local

and state governments, and the public. The unresolved

question is do these wastes require special handling

and disposal, or can they be disposed of by the same

methods used to dispose of conventional, nonradioactive wastes? Presently, the federal government does not

regulate the disposal of naturally occurring radioactive

materials (NORM) waste from drinking water treatment

processes. The U.S. Nuclear Regulatory Commission

(NRC) has the authority to regulate the handling and

disposal of all licensed radioactive materials, but this

does not include NORM in drinking water. Licensed

materials include high-level radioactive wastes, such as

spent nuclear reactor fuel rods and wastes from nuclear

weapons processing centers and low-level radioactive

wastes such as those used in building materials, machinery, and clothing from nuclear power plants, as well

as spent nuclear medicines.

Although federal agencies do not classify NORM wastes

as radioactive wastes, policies could change, and definitions to characterize these wastes as radioactive could

be developed in the future. Because of current concern

for the proper disposal of these wastes, in July 1990

EPA’s Office of Drinking Water published Suggested

Guidelines for the Disposal of Drinking Water Treatment

Wastes Containing Naturally Occurring Radionuclides

(U.S. EPA, 1990c; Parrotta, 1991). These guidelines

were intended to help WTPs and state and local regulators achieve safe and responsible waste management

practices for water treatment residuals containing radionuclides at concentrations in excess of background

levels.

The EPA guidelines clearly state that no federal regulations specifically address the disposal of wastes concentrated by water treatment processes on the basis of

their naturally occurring radioactive content. The document further states that the guidelines developed were

taken from a review of the regulations and from guidelines originating in other industries and programs. For

177
example, one source of information for the suggested

guidelines for radium wastes was disposal criteria for

WTP residuals that the states of Illinois and Wisconsin

(Illinois, 1984; Wisconsin, 1985) developed. Some states

have also established rules for the disposal of wastes

that contain radioactivity. EPA emphasizes in its guidelines, therefore, that where radioactivity in water treatment residuals is anticipated, state agencies should

always be consulted.

A review of the current literature shows that almost all

information on the disposal of WTP residuals containing

radionuclides comes from the EPA guidance document.

The guidelines are being revised, but no major changes

are anticipated. Until formal regulations are developed,

states having no specific rules or regulations will likely

cite and use this document as general policy. Consequently, the sections on waste disposal in this chapter

have been summarized from EPA’s revised 1994 draft

guideline document (U.S. EPA, 1994b). The revised

guidelines will not be issued to the public until the regulations are promulgated, some time after December 1995.

10.2 Waste Disposal Practices

A list of residuals generated from common water treatment practices appears in Table 10-1. Except for disposal at a radioactive waste disposal site, current

disposal methods are basically the same as those used

for nonradioactive wastes. These methods are shown in

Table 10-2.

Federal, state, and local regulations exist that govern

and limit the discharge of nonradioactive WTP residuals

into the environment. These regulations also apply to

residuals containing radionuclides. The radioactivity of

these residuals causes additional concern with disposal and thus may narrow disposal options or place

additional requirements on the options. In selecting the

disposal option, the concentration of radioactive contaminants in the residuals is the governing factor. Unfortunately, no current federally established levels of

radionuclides have been developed to define low or high

radioactive wastes or dictate the acceptable disposal

method.

10.3 Waste Disposal Guidelines

The following information on disposal of liquids, solids,

and sludges has been summarized from EPA’s revised

1994 draft guideline document (U.S. EPA, 1994b).

10.3.1 Liquid Disposal

10.3.1.1 Discharge to Surface Waters

The Clean Water Act requires that dischargers of pollutants

to navigable waters obtain National Pollutant Discharge

Elimination System (NPDES) permits containing, at a

minimum, technology-based effluent limitations that reflect various levels of wastewater treatment and, where

necessary, more stringent limitations necessary to ensure attainment and maintenance of state water quality

standards. EPA has not promulgated any rule establishing technology-based effluent limitations applicable

to WTPs nationwide.

In the absence of such a categorical standard, limitations are established on a case-by-case basis, using the

best professional judgment. After determining minimum

technology-based requirements, the effect(s) of discharge on the receiving water is also determined; if

necessary, more stringent limitations are established to

protect the receiving water quality (see Chapter 5). In

Table 10-1. Summary of Residuals Produced From Water

Treatment Processes

Water Treatment 

Process Waste Streams/Material

Conventional

coagulation/filtration

Filter backwash water

Sludge (alum or iron)

Filter material

Lime softening Filter backwash water

Sludge (lime)

Filter material 

Ion exchange Reuse and backwash water

Regeneration liquid (brine, caustic, acid)

Resin

Reverse

osmosis/electrodialysis

Reject water

Membrane/Material

Greensand filtration Filter backwash water

Sludge

Greensand media

Selective sorbents

(GAC, resins activated

alumina)

Sorbent media

Table 10-2. Water Treatment Methods for Residuals

Containing Radionuclides

Residuals

Type Disposal Method

Liquids Direct discharge to surface water

Direct discharge to sanitary sewer

Deep well injection

Irrigation

Lagooning/Evaporation ponds

Sludge Lagooning (temporary)

Landfill disposal

• No pretreatment

• With prior lagooning

• With mechanical dewatering

Land disposal

Licensed low-level radioactive waste disposal site

Solids Landfill disposal

Licensed low-level radioactive waste disposal facility

178
certain states, water quality standards may include specific criteria for radionuclides. State water quality standards that require more stringent discharge limitations

must be reflected in any NPDES permit.

The NPDES permitting process should include study of

the flow and geometry of receiving storm sewers and

surface waters, as well as the potential uses of the

surface waters (e.g., drinking water, agriculture). Based

on site conditions, a state or other regulatory agency

may use its discretion in determining a limiting concentration (e.g., a drinking water MCL) or a percentage of

the background concentration(s) of radionuclides (e.g.,

10 percent) that would limit the increase of radionuclides

in the water body and/or the sediments due to the

discharge of WTP residuals.

The NRC has established levels of radionuclides that its

licensees are allowed to release into unrestricted areas

of the environment (10 CFR Section 20.1302(b)(2)(i)

and Sections 20.1001-20.2401). The referenced NRC

limits are: 60 pCi/L for Ra-226 and Ra-228 and 300 pCi/L

for natural uranium. If no state or local standards are in

place, the NRC standards may be used as guidelines

for surface water discharge. Some states and local

authorities have promulgated conservative limits of 10

percent of the NRC levels under 10 CFR 20 for release

of radionuclides into the environment.

If the conditions of flow and geometry are not adequate

to prevent a buildup of radionuclides in surface water or

sediments to within the limit set by the regulator, then

other solutions need to be studied. These may include

additional waste treatment, waste storage, and controlled discharge measures to produce a waste stream that

can meet in-stream requirements. Otherwise, discharge

to surface waters should not be allowed, and other

options need to be considered.

10.3.1.2 Discharge Into Sanitary Sewers

Federal regulations prohibit discharges to sewers that

would cause a municipal waste water treatment plant

(WWTP) to violate an NPDES permit or that would

interfere with wastewater treatment operations or sludge

disposal. In addition, states or localities may establish

more stringent limitations on the discharge of wastes

from WTPs into a sanitary sewer, or may require pretreatment of the waste prior to its release into the sanitary sewer. State or local regulations limiting the

discharge of WTP residuals that contain naturally occurring radionuclides into sanitary sewers would govern

those discharges.

The NRC limits the discharge by its licensees of wastes

containing radioactive materials into sanitary sewers.

For NRC licensees, the monthly quantity of soluble Ra226, Ra-228, and natural uranium, diluted by the average monthly quantity of total WTP residuals released

into a sewer, should not exceed 600 pCi/L, 600 pCi/L,

and 3,000 pCi/L, respectively. Also, the gross quantity of

all radioactive material (excluding tritium and carbon-14)

that a facility releases into sanitary sewer should not

exceed 1 curie per year, according to the same NRC

standards. EPA puts forth these standards for consideration in this context.

The NRC standards correspond to an individual radiation dose of 50 millirem per year. Although high, the

reasonable dilution that a receiving waterway provides

may reduce this dose to levels below any applicable

drinking water standards. This should be examined,

however, on a case-by-case basis.

If accumulation of radioactivity in the sanitary sewer

distribution system, sewage treatment facility or publicly

owned treatment works (POTW) is observed, discharging radioactive residuals into a sanitary sewer should be

discontinued until radiation exposures and possible hazards to personnel repairing sewage pipelines are evaluated. Discharge of residuals containing radionuclides

into a sanitary sewer results in the accumulation of

radionuclides in the sludge that the POTW produces.

Affected sewage sludge should be monitored for high

levels of radioactivity and disposed of properly.

10.3.1.3 Deep Well Injection

Under Part C of the Safe Drinking Water Act (SDWA),

EPA is required to promulgate minimum requirements

for effective underground injection control (UIC) programs to prevent endangering underground sources of

drinking water by subsurface emplacement of fluids

through wells. Regulations may be implemented by

states that have adopted requirements at least as stringent as the federal requirements and have been given

primary enforcement responsibility for the UIC program.

For states that do not have primacy, EPA has promulgated state-specific regulations that EPA regional offices

implement.

A WTP interested in discharging residuals containing

radionuclides into an injection well in a primacy state

should first consult with the appropriate state agency.

State regulations may be more stringent than federal

requirements and may actually ban such a practice. The

WTPs in other states should consult with the appropriate

UIC regional branch office of EPA before deciding to

dispose of residuals containing radionuclides into an

injection well in accordance with a method recommended below.

Under the federal requirements, regulation of WTP residuals containing radionuclides depends on the concentrations of radionuclides present. Furthermore,

requirements are specified for shallow and for deep well

injection. Shallow wells are defined as those above or

in an underground source of drinking water (USDW).

179
USDWs and injection wells are defined very broadly. A

USDW is an aquifer or its portion that 1) supplies any

public water system or that contains sufficient quantity

of water to supply a public water system and 2) supplies

drinking water for human consumption, or contains

fewer than 10,000 mg/L TDS. A well is any bored, drilled,

or driven hole where the depth of the hole is greater than

the largest surface dimension. This definition includes

septic systems and cesspools used for disposal of wastes.

Radioactive wastes are treated differently from nonradioactive wastes under the UIC program. As defined in

40 CFR Section 144.3, radioactive waste means any

waste that contains radioactive concentrations that exceed those listed in 10 CFR 20, Appendix B, Table 11,

Column 2. The concentration for Ra-226 and Ra-228 is

currently listed as 60 pCi/L, while the radioactive concentration for natural uranium is 300 pCi/L.

Radioactive wastes as defined by the UIC program (i.e.,

wastes containing greater than 60 pCi/L Ra-226 or Ra228, and/or containing greater than 300 pCi/L uranium)

would not be disposed of in a shallow well, as defined

above. Shallow injection of radioactive wastes (e.g.,

injection above or into a USDW) is a banned practice

under the UIC program. The definitions of USDW and

shallow injection would virtually eliminate any shallow

disposal of radioactive waste.

Well disposal of radioactive waste below a USDW is

currently considered a Class V well injection and is

under study by EPA as part of the Class V regulatory

development effort. At this time, EPA is not prepared to

make any recommendations regarding these wells. The

following is suggested guidance for nonradioactive

wastes as defined by the UIC program (i.e., wastes

containing less than 60 pCi/L Ra-226 or Ra-228, or less

than less than 300 pCi/L uranium).

Well injection of nonradioactive WTP residuals beneath

the lowest USDW is classified as a Class I nonhazardous practice, provided the waste contains no other hazardous components. EPA recommends disposal of

nonradioactive waste through a Class I well because

these wells must be permitted, and current permitting

requirements for these wells ensure adequate protection of USDWs and human health.

10.3.1.4 Other Options

If, due to the properties of a residual, or local regulatory

restrictions, a residual containing NORM cannot be

processed using one of the above methods, then the

treatment operator may choose from another treatment

or disposal option such as evaporation of liquid wastes,

sand drying or lagooning, chemical precipitation of contaminants, or other solids separation techniques. Other

state regulations may apply to these practices. At a

minimum, no practice should be less environmentally

protective than the options mentioned above.

For instance, lagooning of radioactive wastes would be

analogous to shallow well injection, if practiced in an

unlined unit. Lagoons or other impoundments should, at

a minimum, be lined to prevent infiltration. If WTPs

intend to evaporate radioactive waste, the evaporation

unit should be designed and operated properly to ensure

isolation of the waste from the watertable.

10.3.2 Solids and Sludge Disposal

The following disposal guidance for radioactive WTP

residuals is summarized from the EPA guidelines. The

guidance is based solely on the radioactivity of these

wastes and does not address other potentially hazardous substances.

10.3.2.1 Landfill Disposal

Wastes containing less than 3 pCi/g (dry weight) of

radium and less than 50 g/g of uranium may be disposed

of in a municipal landfill without the need for long-term

institutional controls if the wastes are first dewatered

and then spread and mixed with other materials when

they are emplaced. The total contribution of radioactive

wastes to the landfill should constitute only a small

fraction (less than 10 percent of the volume) of the

material in the landfill.

Wastes containing 3 to 50 pCi/g (dry weight) of radium

should be disposed of using a physical barrier (i.e., a

cover) that protects against radon release and isolates

the wastes. Institutional controls designed to avoid inappropriate uses of the disposal site should also be provided. A physical barrier consisting of 10 feet of cover

earth or nonradioactive waste, properly designed for

long-term stability of the waste, should suffice.

Sludge should be dewatered prior to disposal to minimize migration of contaminants. Consideration should

be given to the hydrogeology of the site and other

factors affecting long-term stability of the wastes. Sites

that fully comply with EPA’s Subtitle D regulations and

guidance under the Resource Conservation and Recovery Act (RCRA) would be appropriate for disposal of

these wastes.

A jurisdiction may choose to ensure ground-water protection by specifying RCRA hazardous waste requirements (e.g., properly lined waste unit or sludge

stabilization), to prevent seepage of contaminants from

the landfill. The degree of additional protection a jurisdiction wants to provide against intrusion and misuse

may vary from site to site but should be determined

before waste disposal.

Disposal of solid wastes containing 50 to 2,000 pCi/g

(dry weight) of radium should be determined on a case180
by-case basis. Methods that comply with EPA’s standards for disposal of uranium mill tailings should be considered (40 CFR 192). A decision not to employ such

methods fully should be based on a demonstration of

significant differences between the quantity and potential for migration of uranium mill tailings versus water

treatment residuals. The disposal method should be

augmented by long-term institutional controls to avoid

future misuse of disposal sites; such controls are not

normally already in place at sanitary landfills.

At a minimum, disposal in RCRA-permitted hazardous

waste units should be considered. In states where disposal is licensed or permitted, disposal at these sites

should be considered for radium-bearing solid wastes. At

concentrations approaching 2,000 pCi/g, waste disposal

within a licensed low-level radioactive waste disposal

facility or a facility that is permitted by EPA or a state to

dispose of discrete wastes, should be considered.

In states where lower concentration waste disposal is

licensed or permitted, that option should be considered

for disposal of solids containing 50 to 500 g/g (dry

weight). Recovery of the uranium resource (e.g., at a

uranium milling site) may be considered. NRC may

require licensing of the recovered material (greater than

0.05 percent), however, as a source material under the

provision of the Atomic Energy Act. Disposal at a licensed low-level radioactive waste facility should also

be considered.

Solid wastes containing more than 2,000 pCi/g (dry

weight) of radium or more than 500 %g/g (dry weight)

uranium should be disposed of in a low-level radioactive

waste disposal facility or at a facility that is permitted by

EPA or a state to dispose of NORM wastes. Recovery

of the uranium resource (e.g., at a uranium milling site)

may be considered. NRC may require licensing of the

material containing uranium (if greater than 0.05 percent

uranium) as a source material under the provisions of

the Atomic Energy Act. Also, when radium occurs with

uranium, it is regulated as part of that source material.

10.3.2.2 Land Disposal

EPA does not recommend applying, mixing, or otherwise

spreading NORM wastes onto open land (e.g., farm

land, pasture, orchard or forestry lands, construction

sites, roadbeds) for several reasons:

• Data relating to plant, animal, and human uptake and

the potential exposure that may result from land-applied NORM wastes need to be collected and analyzed. Preliminary assessments suggest that human

health risks from radon inhalation significantly increase in buildings constructed on such areas.

• The long-term control of and monitoring at a site that

may contain higher than background levels of radionuclides (which sometimes are very long-lived)

cannot be assured.

• Diluting wastes runs counter to EPA’s general policy

of concentrating wastes before safe disposal.

• EPA has not collected or reviewed any data on the

status of surface runoff from sites that host land application of WTP residuals. Preliminary risk assessments using conservative assumptions, however,

indicate that runoff and surface water contamination

may pose a significant risk to the general population.

• Although certain types of sludge have been found to

have beneficial properties as amendments to agricultural soils, EPA has not decided that the beneficial

results outweigh the potential adverse results, such

as food-chain contamination, future misuse of sites

for building, and impacts on surface and groundwater quality.

10.4 Recordkeeping

EPA guidelines recommend that water treatment facilities keep records of the amount and composition of

radioactive wastes (solid and liquid) they generate and

the manner and location of their disposal. Repositories

of wastes containing more than 50 pCi/g (dry weight)

should be permanently marked to ensure long-term protection against future misuse of the site and/or its materials. This guideline is not meant to cause an increase

in the federal recording requirements for water treatment facilities.

181
Chapter 11

Economics

Any water utility manager considering a new residuals

management option for a water treatment plant (WTP)

will want to estimate as closely as possible the cost of

the residuals management option. This chapter provides

cost equations that can be used to estimate the total

annual costs associated with several different residuals

treatment and disposal options. The cost equations provided here were developed using computer cost models, published cost information, and quotes provided by

vendors. All costs presented in this chapter are in 1992

dollars.

These equations should be used for relative comparison

purposes only. Certainly, more accurate, site-specific

estimates should be developed for facility planning and

budgeting purposes. The specific factors and treatment

requirements will likely drive the costs up.

Final project costs have been found to be higher than

those costs presented herein. This is primarily due to

site-specific factors such as physical and chemical

characteristics of the residuals, site constraints, operational costs, and differing state and local regulatory

requirements.

11.1 Cost Assumptions

11.1.1 Capital Cost Assumptions

Table 11-1 presents cost factors and unit costs used to

calculate the capital costs presented in this chapter. The

land, buildings, and piping components are discussed in

detail below. Other capital cost items (e.g., tanks, lagoons, and pumps) are calculated individually.

11.1.1.1 Land

Land prices vary considerably across the country and

depend on proximity to metropolitan areas, the state of

land development (i.e., improved or unimproved), current use, and scarcity of land. A cost of $10,000 per acre

is assumed for the cost estimates included in this chapter. Costs for suburban or industrial unimproved land

average $10,000 per acre and range from approximately $4,000 to $350,000 per acre depending on location. Costs for unimproved agricultural or rural land vary

from approximately $150 per acre in states with large

tracts of undeveloped land, to $2,200 per acre in states

with small tracts of undeveloped land. An average cost

for rural, unimproved land is $1,000 per acre.

Costs for purchasing land are included in the capital cost

equations provided here for all treatment and disposal

processes that use surface impoundments. For example, assume an impoundment is used for liquid residuals

storage prior to land application. The land costs are

presented as a separate cost in the capital cost equation

to allow the user to delete the cost for land if no purchase

is required or to vary the cost of the land if the unit cost

of land for a given area is available. A minimum land

purchase of a 1/2 acre is used in developing land costs.

Land costs are not included for direct discharge and

discharge to a publicly owned treatment works (POTW).

It is assumed that the water treatment authority is able

to obtain an easement for placement of a sewer or

discharge line.

11.1.1.2 Buildings

Building costs also vary considerably based on geographic location. The building cost of $33.00 per square

foot, listed in Table 11-1, comes from 1992 Means Building Construction Cost Data (R.S. Means Co., Inc.,

Table 11-1. Capital Cost Factors and Selected Unit Costs for

WTP Facility Planning (U.S. EPA, 1993c)

Component Factor/Unit Cost

Land $10,000/acre

Buildings $33.00/ft2

Piping 5% of installed equipmenta

Pipe fittings 20% of piping costsb

Electrical 1% of installed equipment

Instrumentation 1–2% of installed equipment

Engineering fee 15% of direct capital

Contingency, bonding, and

mobilization

20% of direct capital

Contractor’s overhead and profit 12% of direct capital

a Piping costs are calculated directly when piping is a significant cost

(e.g., for direct discharge). b Factor is used when piping costs are calculated directly.

182
1991). This cost is for a basic warehouse or storage

building and includes electrical, foundations, heating,

ventilation, air conditioning, and plumbing costs. It includes a 10 percent surcharge for structures of less than

25,000 square feet. This cost is the median cost for

basic warehouse and storage buildings. The lower quartile cost is $21.30 per square foot and the upper quartile

cost is $45.65 per square foot. Different warehouse or

building configurations will change this unit cost. An

aesthetically pleasing architecturally designed building

or a complex building with several levels may cost substantially more than the building costs included in this

chapter.

11.1.1.3 Piping and Pipe Fittings

Piping costs are calculated using two different methodologies that depend on the magnitude of the piping

costs. For management systems in which the piping

constitutes a significant portion of the capital cost (e.g.,

direct discharge and discharge to a POTW), piping costs

are calculated by sizing individual pipe lengths and determining installation charges. For these management

systems, pipe fittings are assumed to be 20 percent of

the installed piping costs. For management systems

where piping is not a significant cost (e.g., chemical

precipitation and mechanical dewatering), piping and

pipe fittings are assumed to be 5 percent of the installed

equipment costs.

11.1.2 Operation and Maintenance

Assumptions

Table 11-2 presents cost factors and unit costs used to

calculate the operation and maintenance costs presented in this chapter.

11.1.2.1 Labor

The labor rate is based on a 20-city average for union

scale laborers. The hourly rate of $28.00 is based on an

average salary of $18.74 per hour including fringe benefits and a 50 percent labor overhead cost. Average

salaries, including fringe benefits, by region range from

$11.17 per hour in the southeast to $26.23 in the western United States.

11.1.2.2 Supervision

The supervision rate is based on an estimated salary of

$45,000 per year for a water system treatment manager.

The $42.00 per hour assumes a 30 percent fringe benefit and a 50 percent labor overhead cost. The supervisor’s salary is based on 1992 job postings in Public

Works magazine (Public Works, 1992).

11.1.3 Total Annual Cost Assumptions

Total annual costs are equal to the sum of the yearly

operation and maintenance costs plus the annualized

capital costs. Annualized capital costs are calculated

based on a 20-year operating life and an interest rate of

10 percent. The capital recovery factor for a 20-year

operating life at a 10 percent interest rate is 0.1175.

Alternate capital recovery factors can be calculated using the formula presented below.

Capital recovery factor = (1 + i)

N (i)

(1 + i)

N − 1

where

i = interest rate

N = number of years

11.1.4 Cost Components Excluded

The following cost components are not included in the

cost estimates presented in this chapter. Other excluded

costs, specific to each management option, are highlighted under their respective technology cost sections.

11.1.4.1 Sample Collection and Laboratory

Analysis

Costs for sample collection and analysis to determine

solids content, free liquids, toxicity characteristics, and

other parameters are not included in the costs presented in

this chapter. Sampling frequency is highly variable. Depending on the residuals management method selected,

sampling requirements could be minimal or extensive.

11.1.4.2 Permits and Other Regulatory

Requirements

Costs for permits and other regulatory requirements are

not included in this chapter. Requirements vary considerably from state to state for a given management option. Permitting costs will vary based on the size and

complexity of a unit and the local governing jurisdiction.

Management methods that may require permits include

landfills, land application, evaporation ponds, and storage lagoons. In addition, generators of hazardous waste

Table 11-2. Operation and Maintenance Cost Factors and

Unit Costs for WTP Facility Planning (U.S. EPA,

1993c)

Component Factor/Unit Cost

Labor $28.00/hour

Supervision $42.00/hour

Supervision factor 10% of labor hours

Insurance and general and

administrative expenses

2% of direct capital, excluding

land and buildings

Maintenance 2 to 3% of direct capital,

excluding land and buildings

Electricity $0.086/kilowatt-hour

183
are required to comply with RCRA generator requirements (40 CFR Part 262).

11.1.4.3 Mechanical Standby Equipment

Costs for standby equipment, or redundancy, are not

included in these cost estimates.

11.1.5 Cost Equations

The cost equations for the capital components and for

operation and maintenance were developed by estimating the costs for different residuals flow rates. Equipment sizes, retention times, installed equipment costs,

operation and maintenance expenses, commercial disposal fees, component cost factors, and unit costs were

developed based on published literature, (e.g., AWWA

reports, technical journal articles, EPA documents), vendor contacts, and best engineering judgment. Capacities for major equipment components are presented in

each cost section. Each cost equation covers the range

of expected residuals flow rates unless a specific technology is not applicable to the entire range. For example, some technologies are not feasible for very high

flow rates because of excessive land requirements or

equipment limitations.

The limits of the cost equations presented for each

residuals management method may be extended

slightly beyond the ranges presented in this chapter. The

upper or lower end for some equations, however, may

represent the largest or smallest practical limits of a

specific technology. Consequently, the design assumptions should be carefully reviewed before extending the

range of the equations. Best engineering judgment

should be used when extending the equations beyond

their specified limits.

The capital and operation and maintenance equations

encompass the residuals volumes generated by WTPs

ranging from a design flow rate of 1.8 million gallons per

day (mgd) (average operating flow rate, 0.7 mgd) to 430

mgd (average operating flow rate, 270 mgd). One option

is to size the equipment (i.e., determine the capital cost)

for the residuals produced from the water treatment

system design flow rate, while calculating the operating

cost using the average residuals flow rate. Thus, the

equipment is sized for peak flow periods, but the actual

operating costs may be more closely estimated using an

average daily flow rate.

Typically, the design flow rate for a WTP may be two to

three times the average flow rate for the facility. When

calculating the costs for capital intensive management

technologies such as chemical precipitation and mechanical dewatering, calculate both the capital and operation and maintenance costs using the average flow

rate. Table 11-3 indicates whether the water treatment

design flow rate or average flow rate should be used to

calculate the capital and operation and maintenance

costs. The design flow rate can be used to calculate

costs, but this calculation may result in substantial overestimation of the cost to treat WTP residuals generated

by an actual facility.

All costs presented in this chapter are in 1992 dollars.

To update cost results obtained from either the cost

equations or cost curves specified in the following sections, the Builders’ Construction Cost Indexes are recommended and available in quarterly “cost report”

issues of Engineering News Record (ENR). The ENR

20-city construction cost index or the Means Construction Cost Index are good general-purpose cost updating

indexes. Their January 1992 values are 455.08 and

97.90, respectively.

11.1.6 Cost Curves

The cost curves in this chapter are the direct graphical

representations of the cost equations. A break in the

curve indicates that two equations covered two different

ranges of the data. In these instances, the two curves

typically do not meet at the same point at the overlap.

11.1.7 Calculating Residuals Management

Costs

WTP residuals management costs are calculated using

the capital, operating and maintenance and total annual

cost equations presented in the following sections.

Known residual flow rates from an operating facility may

be used or residuals volumes can be estimated using

the information presented in Chapter 3. The annual

residuals management costs can be determined by calculating a total annual cost for each required treatment

or disposal technology and then summing the costs

Table 11-3. Flow Rate Use in Calculating Facility Costs (U.S.

EPA, 1993c)

Technology

Capital

Cost

Operation &

Maintenance Cost

Gravity thickening Average Average

Chemical precipitation Average Average

Mechanical dewatering Average Average

Nonmechanical dewatering Design Average

Evaporation ponds Design Average

POTW discharge Design Average

Direct discharge Design Average

French drain Design Average

Land application Design Average

Nonhazardous waste disposal Average Average

Hazardous waste disposal N/A N/A

Radioactive waste disposal N/A N/A

184
associated with all WTP residuals management processes at the facility.

11.2 Gravity Thickening

The following section presents the design assumptions,

capital components, and the operation and maintenance

components used to estimate the costs for gravity thickening of filter backwash streams from chemical coagulation, or lime softening residuals streams and/or their

sludge waste streams.

11.2.1 Design Assumptions

• Gravity thickening is used to pretreat filter backwash

or sludges from chemical coagulation and lime softening.

• The waste streams flow by gravity from the treatment

plant to settling tank.

• The backwash volume ranges from 7,000 to 2,500,000

gallons per day.

• The total backwash volume is 0.5 to 2 percent of the

design flow rate.

• The initial solids concentration of the backwash is 0.1

percent and the discharge concentration is equal to

1 percent.

• The backwash is thickened in a tank.

• The volume of backwash is reduced by 90 percent.

• The supernatant is pumped to the head of the treatment plant.

• The thickened sludges are discharged to another

treatment for further dewatering.

11.2.2 Capital Components

The capital components for each filter backwash system

consist of the following items: holding tank, piping and

fittings, pump, trenching, electrical, and instrumentation.

Table 11-4 indicates the holding tank capacities used to

develop the capital cost equation for gravity thickening.

11.2.3 Operation and Maintenance

Components

The operation and maintenance components for each

gravity thickening system include: electricity, labor and

supervision, maintenance labor and materials, insurance, and general and administration.

11.2.4 Cost Components Excluded

The following cost components are not included in the

cost equations for gravity thickening systems:

• Supernatant disposal, if the supernatant cannot be

pumped to the head of the treatment plant.

• Thickened sludge disposal.

• Costs for supernatant disposal, if the supernatant

cannot be pumped to the head of the treatment plant,

and treatment of the thickened sludge can be determined using other sections of this chapter.

11.2.5 Gravity Thickening Cost Equations

and Cost Curves

The cost equations for the capital components and the

operation and maintenance components for gravity

thickening were developed by estimating the costs for

nine different filter backwash flow rates. The filter backwash and sludge flow rates (X) used in the equations

below are the average daily volumes generated (e.g.,

for filter backwash, 2.5 percent of the daily treated water

flow rate). The capital cost equation calculates the total

capital cost (i.e., installed capital plus indirect capital

costs). The total annual cost is calculated based on the

capital and operation and maintenance costs obtained

using the equations presented below. Capital costs for

gravity thickening are presented graphically in Figure

11-1 and operation and maintenance costs are shown

in Figure 11-2.

Capital Costs

Y = 32,400 + [41.7(X0.50)]

where

Y=$

X = gallons of filter backwash/sludge per day

Range: 7,000 gpd ≤ X ≤ 270,000 gpd

Y = 21,700 + [66.5(X0.50)]

where

Y=$

X = gallons of filter backwash/sludge per day

Table 11-4. Holding Tank Capacities, Gravity Thickening (U.S.

EPA, 1993c)

Waste Stream

Volume (gpd)

Number of

Tanks

Waste Stream Settling

Tank Capacity (gal)

 7,000 1 30,000

 13,300 1 50,000

 36,000 1 30,000

 96,000 1 30,000

 165,000 1 50,000

 270,000 1 50,000

 510,000 1 50,000

1,050,000 2 50,000

2,150,000 2 50,000

185
Range: 270,000 gpd < X ≤ 2,500,000 gpd

Operation and Maintenance

Y = 5,800 + [8.04(X0.50)]

where

Y = $/year

X = gallons of filter backwash/sludge per day

Range: 7,000 gpd ≤ X ≤ 2,500,000 gpd

Total Annual Cost

Y = [CAP(CRF)] + O&M

where

Y = $/year

CAP = capital cost

CRF = capital recovery factor (e.g., 0.1175)

O&M = operation and maintenance cost

11.3 Chemical Thickening

This section presents the design assumptions, capital

components, and the operation and maintenance components used to estimate the costs for chemical thickening of residuals streams such as ion exchange

backwash and reverse osmosis brines.

11.3.1 Design Assumptions

• The chemical precipitation system consists of mixing

and holding tanks for the lime solution, a precipitation

tank, a clarifier, agitators, and pumps.

• The precipitation tank has a 1/2-hour retention time

and a 5 percent overdesign.

• The clarifier (settling tank) has a 1- to 2-hour retention

time.

• Waste brines flow to the treatment system under

pressure from the ion exchange or reverse osmosis

system.

• The chemical precipitation equipment is located in

the water treatment building.

• Sludge from the clarifier may require additional dewatering prior to disposal.

• The sludge volumes generated by chemical precipitation are 5 and 2 percent of the influent volumes for

reverse osmosis and ion exchange brines, respectively.

11.3.2 Capital Components

The capital components for each chemical precipitation

system consist of: carbon steel mixing tank, carbon steel

holding tank, carbon steel precipitation tank, clarifier,

agitators, sludge pumps, building, piping, electrical and

instrumentation.

Table 11-5 indicates the brine volumes, tank capacities,

and clarifier capacities used to develop the capital cost

equation for chemical precipitation.

Figure 11-1. Capital costs for gravity thickening. Figure 11-2. O&M costs for gravity thickening.

186
11.3.3 Operation and Maintenance

Components

The operation and maintenance components for each

chemical precipitation system include: lime, electricity,

labor, maintenance labor and materials, insurance, general and administration, and water.

11.3.4 Cost Components Excluded

The following cost components are not included in the

cost equations for chemical precipitation systems since

they are accounted for elsewhere:

• Sludge dewatering (if needed for the selected disposal option).

• Sludge disposal.

• Clarifier overflow disposal, if the overflow cannot be

pumped to the head of the treatment works.

Costs for sludge dewatering and sludge disposal techniques are included in other sections of this chapter.

Costs for clarifier overflow disposal, if the overflow cannot be pumped to the head of the treatment works, can

be determined using other sections of this chapter.

11.3.5 Chemical Precipitation Cost

Equations and Cost Curves

The cost equations for the capital components and operation and maintenance were developed by estimating

the costs for six different brine flow rates. The capital

cost equation calculates the total capital cost (i.e., installed capital plus indirect capital costs). Equipment

costs and building costs are separate components in the

capital cost equation to enable the user to exclude the

cost for a building addition if it is not necessary. The total

annual cost is calculated based on the capital and operation and maintenance costs obtained using the equations presented below.

A capital cost curve for chemical precipitation is shown

in Figure 11-3. Operation and maintenance costs for

chemical precipitation are shown in Figure 11-4.

Table 11-5. Capital Cost Equation Determinants, Chemical

Precipitation (U.S. EPA, 1993c)

Residuals

Flow Rate

(gpd)

Precipitation

Tank Size

(gal)

Mix Tank

Capacity

(gal)

Clarifier

Capacity

(gal)

 27,400 650 50 2,500

 67,500 2,000 75 8,000

 274,000 6,000 200 24,000

 500,000 12,000 600 40,000

1,000,000 23,000 1,200 50,000

2,740,000 60,000 1,500 40,000 x 2

Figure 11-3. Capital costs for chemical precipitation.

Figure 11-4. O&M costs for chemical precipitation.

187
Capital Costs

Y = [34,600 + 206(X0.50)] + [110(X0.55)]

Y = [equipment cost] + [building cost]

where

Y=$

X = gallons of brine per day

Range: 25,000 gpd ≤ X ≤ 500,000 gpd

Y = [34,600 + 206(X0.50)] + [65,000 + 160(X0.50)]

Y = [equipment cost] + [building cost]

where

Y=$

X = gallons of brine per day

Range: 500,000 gpd < X ≤ 3,000,000 gpd

Operation and Maintenance

Y = 37,300 + 0.17(X)

where

Y = $/year

X = gallons of brine per day

Range: 25,000 gpd ≤ X ≤ 500,000 gpd

Y = 57,500 + 0.13(X)

where

Y = $/year

X = gallons of brine per day

Range: 500,000 gpd < X ≤ 3,000,000 gpd

Total Annual Cost

Y = [CAP(CRF)] + O&M

where

Y = $/year

CAP = capital cost

CRF = capital recovery factor (e.g., 0.1175)

O&M = operation and maintenance cost

11.4 Mechanical Sludge Dewatering

11.4.1 Pressure Filter Press Cost

Components

This section presents the design assumptions, capital

components, and the operation and maintenance components used to estimate the costs for a pressure filter, or

plate and frame, press for mechanical sludge dewatering.

11.4.1.1 Design Assumptions

• Pressure filter presses are effective for residuals flow

rates greater than 100 gallons per day.

• The polymer feed system consists of a polymer storage tank, a polymer pump, a polymer/sludge contact

tank, and a 20- to 30-minute holding tank.

• A positive displacement pump delivers residuals to

the filter press.

• The filter press operates on a batch basis.

• Filtrate collects in a filtrate holding tank before being

pumped to the head of the treatment plant.

• In smaller systems, the filter cake collects in a small,

wheeled container, which is emptied into a larger

solids bin as necessary.

• In larger systems, the filter cake drops into rolloff bins

located beneath the filter press.

• Accumulated solids require disposal on a periodic

basis.

• The dewatered residuals volume is 0.03 and 10 percent of the initial volume for alum and lime residuals,

respectively.

11.4.1.2 Capital Components

The capital components for each pressure filter press

system consist of the following items: Schedule 40 steel

piping, polymer feed system, positive displacement

pump(s), pressure filter press(es), steel filtrate tank and

pump, filter cake storage bin, building, piping, electrical,

and instrumentation.

Table 11-6 indicates the sludge volumes and filter press

capacities used to develop the capital cost equation for

pressure filter presses.

11.4.1.3 Operation and Maintenance

Components

The operation and maintenance components for each

pressure filter press system include: electricity, polymer,

labor, maintenance labor and materials, insurance, and

general and administration.

11.4.1.4 Cost Components Excluded

The following cost components are not included in the

cost equations for pressure filter press systems:

• Dewatered residuals disposal.

• Filtrate disposal, if the filtrate cannot be pumped to

the head of the treatment plant.

• Acid wash system, if required.

Costs for residuals disposal techniques are included in

other sections of this chapter. Costs for filtrate disposal,

if the overflow cannot be pumped to the head of the

treatment works, can be determined using other sections of this chapter.

188
11.4.1.5 Pressure Filter Press Cost Equations

and Cost Curves

The cost equations for the capital components and operation and maintenance were developed by estimating

the costs for six different residuals flow rates. The capital

cost equation calculates the total capital cost (i.e., installed capital plus indirect capital costs). Equipment

costs and building costs are separate components in the

capital cost equation to enable the user to exclude the

cost for a building addition if it is not necessary. The total

annual cost is calculated based on the capital and operation and maintenance costs obtained using the equations presented below. A capital cost curve for pressure

filter press is shown in Figure 11-5. An operation and

maintenance cost curve is shown in Figure 11-6.

Capital Cost

Y = [15,600 + 1,520(X0.50)] + [4,600 + 3.0(X)]

Y = [equipment cost] + [building cost]

where

Y=$

X = gallons of sludge per day

Range: 500 gpd ≤ X ≤ 50,000 gpd

Y = [4,181(X0.50) − 579,900] + [22,300 + 

(9.8 x 10-2)(X)]

Y = [equipment cost] + [building cost]

where

Y=$

X = gallons of sludge per day

Range: 50,000 gpd < X ≤ 600,000 gpd

Operation and Maintenance

Y = 13,400 + 0.92(X)

where

Y = $/year

X = gallons of sludge per day

Range: 500 gpd ≤ X ≤ 50,000 gpd

Y = 32,900 + 0.67(X)

where

Y = $/year

X = gallons of sludge per day

Range: 50,000 gpd < X ≤ 600,000 gpd

Table 11-6. Capital Cost Equation Determinants, Pressure

Filter Presses (U.S. EPA, 1993c)

Sludge Flow Rate

(gpd)

Filter Press Capacity

(ft3

)

 500 2.0

 2,000 10

 5,000 24

 50,000 50

250,000 130 x 2

600,000 160 x 3

Figure 11-6. O&M costs for pressure filter press.

Figure 11-5. Capital costs for pressure filter press.

189
Total Annual Cost

Y = [CAP(CRF)] + O&M

where

Y = $/year

CAP = capital cost

CRF = capital recovery factor (e.g., 0.1175)

O& = operation and maintenance cost

11.4.2 Scroll Centrifuge Cost Components

This section presents the design assumptions, capital

components, and the operation and maintenance

components used to prepare the costs for the solidbowl, or scroll centrifuge, used for mechanical

sludge dewatering.

11.4.2.1 Design Assumptions

• The polymer feed system consists of polymer storage

tank, a polymer pump, and a polymer/sludge contact

tank, and a 20- to 30-minute holding tank.

• A positive displacement pump feeds conditioned

sludge to the scroll centrifuge.

• The scroll centrifuge operates 8 to 24 hours per day,

depending on the size of the water system.

• Centrate collects in a centrate holding tank before

being pumped to the head of the treatment works.

• Dewatered residuals collect in a small, wheeled container that is emptied into a larger solids bin as necessary.

• Large systems are located on the second floor and

the dewatered residuals drop into sludge containers

located below the centrifuge.

• Accumulated solids require disposal on a periodic

basis.

• The dewatered residuals volume is 0.04 of the initial

residuals volume.

11.4.2.2 Capital Components

The capital components for each scroll centrifuge system are: Schedule 40 steel piping, polymer feed system,

positive displacement pump, scroll centrifuge with

backdrive, steel centrate tank and pump, dewatered

sludge storage bin, building, piping, electrical, and

instrumentation.

Table 11-7 indicates the sludge flow rates and centrifuge

motor sizes used to develop the capital cost equation for

the scroll centrifuge system.

11.4.2.3 Operation and Maintenance

Components

The operation and maintenance components for each

scroll centrifuge system include: electricity, polymer, labor, maintenance labor and materials, insurance, and

general and administration.

11.4.2.4 Cost Components Excluded

The following cost components are not included in the

cost equations for scroll centrifuge systems: scroll

cleaning apparatus, dewatered residuals disposal, and

centrifuge overflow disposal, if the overflow cannot be

pumped to the head of the treatment plant.

Costs for residuals disposal techniques are included in

other sections of this chapter. Costs for centrifuge overflow disposal, if the overflow cannot be pumped to the

head of the treatment works, can be determined using

other sections of this chapter.

11.4.2.5 Scroll Centrifuge Cost Equations and

Cost Curves

The cost equations for the capital components and operation and maintenance were developed by estimating

the costs for seven different sludge flow rates. The

capital cost equation calculates the total capital cost

(i.e., installed capital plus indirect capital costs). Equipment costs and building costs are separate components

in the capital cost equation to enable the user to exclude

the cost for a building addition if it is not necessary. The

total annual cost is calculated based on the capital and

operation and maintenance costs obtained using the

equations presented below. A capital cost curve for scroll

centrifuge is shown in Figure 11-7 and operation and

maintenance costs are shown in Figure 11-8.

Table 11-7. Capital Cost Equation Determinants, Scroll

Centrifuge (U.S. EPA, 1993c)

Sludge Flow Rate

(gpd) Centrifuge Motor Size

 2,000 7 1/2 hp

 5,000 7 1/2 hp

 50,000 10 hp

 250,000 25 hp

 600,000 40 hp

2,500,000 80 hp x 2

5,200,000 80 hp x 4

190
Capital Cost

Y = [138,400 + 666(X0.50)] + [15,400 + 59(X0.50)]

Y = [equipment cost] + [building cost]

where

Y=$

X = gallons of sludge per day

Range: 2,000 gpd ≤ X < 250,000 gpd

Y = [287,000 + 0.63(X)] + [40,400 + (2.26 x 10-2)(X)]

Y = [equipment cost] + [building cost]

where

Y=$

X = gallons of sludge per day

Range: 250,000 gpd ≤ X ≤ 5,500,000 gpd

Operation and Maintenance

Y = 17,500 + 311(X0.50)

where

Y = $/year

X = gallons of sludge per day

Range: 2,000 gpd ≤ X < 250,000 gpd

Y = 72,000 + 0.43(X)

where

Y = $/year

X = gallons of sludge per day

Range: 250,000 gpd ≤ X ≤ 5,500,000 gpd

Total Annual Cost

Y = [CAP(CRF)] + O&M

where

Y = $/year

CAP = capital cost

CRF = capital recovery factor (e.g., 0.1175)

O&M = operation and maintenance cost

11.5 Nonmechanical Sludge Dewatering

11.5.1 Storage Lagoons Cost Components

This section presents the design assumptions, capital

components, and the operation and maintenance components used to estimate the costs for residuals dewatering using storage lagoons.

11.5.1.1 Design Assumptions

• Storage lagoons are used to dewater alum residuals

and are periodically dredged.

• Permanent lagoons are used to dewater lime residuals; each storage lagoon has a 10-year capacity.

• The lagoons are earthen basins lined with a synthetic

membrane and a geotextile support fabric. They have

no underdrains, but have a variable height outlet

structure to discharge supernatant. Permanent lagoons are 6- to 12-feet deep and storage lagoons

range from 3- to 5-feet deep.

Figure 11-7. Capital costs for scroll centrifuge. Figure 11-8. O&M costs for scroll centrifuge.

191
• Decanted supernatant from the lagoon is collected

and pumped to the head of the treatment plant.

• Residuals flow from the treatment plant to the lagoon

by gravity via 1,250 to 4,000 feet of 2- to 6-inch

diameter PVC piping. Pipes are laid 4 feet below

grade.

• Alum treatment plants operate two lagoons. Each

lagoon has a 6-month storage capacity. After 6

months, the second lagoon is used while free liquid

from the idle lagoon is allowed to evaporate. The

remaining free liquid is decanted and returned to the

head of the treatment plant. Sludge is dredged from

the lagoon on an annual basis.

• Lime treatment plants operate two or four lagoons to

provide a total of 20 years of permanent storage.

Each lagoon has a 6- or 12-month storage capacity

prior to drying and decanting. Free liquid is continuously decanted from the lagoon while it is accepting

sludge and returned to the head of the treatment

plant. After the 6- or 12-month period, another lagoon

is used while free liquid from the idle lagoon is decanted and evaporated. The remaining sludge is allowed to dry prior to the next sludge application.

• Approximately 1.5 cubic feet of solids per 1,000 gallons of initial alum residuals volume will require final

disposal from the storage lagoons.

11.5.1.2 Capital Components

The capital components for each storage lagoon consist

of: two or four lined earthen lagoons, piping and fittings,

trenching, decant collection pump, electrical, instrumentation, and land clearing.

Tables 11-8 and 11-9 indicate the sludge volumes, number of ponds, and total lagoon surface area used to

develop the capital cost equation for lime softening and

alum storage lagoons, respectively.

11.5.1.3 Operation and Maintenance

Components

The operation and maintenance components for each

storage lagoon include: labor and supervision, maintenance labor and materials, electricity, insurance, general and administration, and annual sludge removal

(alum sludge only).

11.5.1.4 Cost Components Excluded

The following cost components are not included in the

cost equations for storage lagoons: sludge dewatering

equipment, dewatered sludge disposal, and supernatant

disposal, if the supernatant cannot be pumped to the

head of the treatment plant.

Costs for sludge disposal techniques are included in

other sections of this chapter. Costs for supernatant

disposal, if the supernatant cannot be pumped to the

head of the treatment plant, can be determined using

other sections of this chapter.

11.5.1.5 Lime Softening Storage Lagoon Cost

Equations and Cost Curves

The cost equations for the capital components and operation and maintenance were developed by estimating

the costs for six different sludge flow rates. The capital

cost equation calculates the total capital cost (i.e., installed capital plus indirect capital costs). Equipment

costs and land costs are separate components in the

capital cost equation to enable the user to exclude the

cost for purchasing land if it is not necessary. The total

annual cost is calculated from the capital and operation

and maintenance costs obtained using the equations

presented below. A capital cost curve for lime softening

storage lagoon is shown in Figure 11-9. An operation

and maintenance cost curve is shown in Figure 11-10.

Table 11-8. Capital Cost Equation Determinants, Lime

Softening Storage Lagoons (U.S. EPA, 1993c)

Lime Softening

Sludge Volume

(gpd)

Number

of Ponds

Total Lagoon

Surface Area

(acres)

 7,000 2 5

 20,000 2 7

 50,000 2 18

 100,000 2 37

 500,000 4 112

1,000,000 4 224

Table 11-9. Capital Cost Equation Determinants, Alum

Storage Lagoons (U.S. EPA, 1993c)

Alum Sludge Volume

(gpd)

Number

of Ponds

Total Surface

Area (acres)

 250 2 0.1

 1,000 2 0.4

 5,000 2 2 

 10,000 2 4 

 50,000 2 19

100,000 2 22

200,000 2 45

500,000 2 112 

192
Capital Costs

Y = [144(X0.91)] + {[(1.57 x 10-3)(X0.87)](Z)}

Y = [equipment cost] + [land cost]

where

Y=$

X = gallons of sludge per day

Z = land cost in $/acre (e.g., $10,000/acre)

Range: 2,800 gpd ≤ X ≤ 1,000,000 gpd

Operation and Maintenance

Y = 7,700 + 1.69(X)

where

Y=$

X = gallons of sludge per day

Range: 2,800 gpd ≤ X ≤ 50,000 gpd

Y = 33,100 + 1.33(X)

where

Y = $/year

X = gallons of sludge per day

Range: 50,000 gpd < X ≤ 1,000,000 gpd

Total Annual Cost

Y = [CAP(CRF)] + O&M

where

Y = $/year

CAP = capital cost

CRF = capital recovery factor (e.g., 0.1175)

O&M = operation and maintenance cost

11.5.1.6 Alum Sludge Storage Lagoons Cost

Equations and Cost Curves

The cost equations for the capital components and operation and maintenance were developed by estimating

the costs for eight different sludge flow rates. The capital

cost equation calculates the total capital cost (i.e., installed capital plus indirect capital costs). Equipment

costs and land costs are separate components in the

capital cost equation to enable the user to exclude the

cost for purchasing land if it is not necessary. The total

annual cost is calculated from the capital and operation

and maintenance costs obtained using the equations

presented below. A capital cost curve for alum sludge

storage lagoons is shown in Figure 11-11. An operation

and maintenance cost curve is shown in Figure 11-12.

Capital Costs

Y = [13,700 + 60.8(X)] + {[0.5 + (3.44 x 10-4)(X)](Z)}

Y = [equipment cost] + [land cost]

where

Y=$

X = gallons of sludge per day

Z = land cost in $/acre (e.g., $10,000/acre)

Range: 250 gpd ≤ X ≤ 10,000 gpd

Y = [326,500 + 28.6(X)] + {[4.2 + (2.35 x 10-4)(X)](Z)}

Y = [equipment cost] + [land cost]

where

Y=$

Figure 11-9. Capital costs for lime softening storage lagoon. Figure 11-10. O&M costs for lime softening storage lagoon.

193
X = gallons of sludge per day

Z = land cost in $/acre (e.g., $10,000/acre)

Range: 10,000 gpd ≤ X ≤ 500,000 gpd

Operation and Maintenance

Y = 6,100 + 2.56(X)

where

Y = $/year

X = gallons of sludge per day

Range: 250 gpd ≤ X ≤ 10,000 gpd

Y = 28,500 + 1.02(X)

where

Y = $/year

X = gallons of sludge per day

Range: 10,000 gpd ≤ X ≤ 500,000 gpd

Total Annual Cost

Y = [CAP(CRF)] + O&M

where

Y = $/year

CAP = capital cost

CRF = capital recovery factor (e.g., 0.1175)

O&M = operation and maintenance cost

11.5.2 Evaporation Ponds Cost Components

The following section presents the design assumptions,

capital components, and the operation and maintenance

components used to prepare the costs for evaporation

ponds.

11.5.2.1 Design Assumptions

• The evaporation pond is used for treatment of IX or

RO brines. The influent solids concentration ranges

from 1.5 to 3.5 percent by weight.

• RO systems with flow rates of 4.8 mgd and greater,

and IX systems with flow rates of 51 mgd and greater

are not considered applicable for evaporation ponds

since over 150 acres is required to evaporate the

daily generation volumes.

• Waste brines flow from the treatment plant to the

evaporation pond by gravity or by the pressure from

the treatment system via 1,250 to 4,000 feet of 4- to

8-inch diameter PVC piping. Pipes are laid 4 feet

below grade.

• Waste brine flow rates range from 31,000 to 500,000

gallons per day.

• The pond is designed for a geographical region with

a net annual evaporation rate of at least 45 inches

per year.

• The pond has no outlet.

• The pond sides and earthen berm are engineered to

have 2.5 to 1 side slopes and 2 feet of free board.

• Soils cut from the excavation of the basin are used

to construct the berm.

Figure 11-11. Capital costs for alum sludge storage lagoon. Figure 11-12. O&M costs for alum sludge storage lagoon.

194
• The pond is constructed with a synthetic membrane

liner and a geotextile support fabric; 1 foot of sand is

placed on top of the liner.

• Piping from the treatment process is sized to discharge the waste brines to the ponds as they are

generated. The treatment process is assumed to run

8 to 16 hours per day.

• The evaporation ponds are sized with sufficient surface area to evaporate the average daily flow. The

pond depth is 2 feet, which provides solids storage

volume and accommodates peak flows.

11.5.2.2 Capital Components

The capital components for each evaporation pond are:

evaporation ponds, piping and fittings, pumps, land

clearing, and instrumentation.

Table 11-10 indicates the brine volumes, number of

ponds, and the total required surface area used to develop the capital cost equation for evaporation ponds.

11.5.2.3 Operation and Maintenance

Components

The operation and maintenance components for each

evaporation pond include: mowing, labor and supervision, maintenance labor and materials, insurance, and

general and administration.

11.5.2.4 Cost Components Excluded

The cost component not included in the cost equations

for evaporation ponds is dewatered sludge disposal.

Costs for sludge disposal are included in other sections

of this chapter.

11.5.2.5 Evaporation Ponds Cost Equations and

Cost Curves

The cost equations for the capital components and for

operation and maintenance were developed by estimating the costs for six different brine flow rates. The capital

cost equation calculates the total capital cost (i.e., installed capital plus indirect capital cost). The total annual

cost is calculated from the capital and operation and

maintenance costs obtained using the equations presented below. A capital cost curve for evaporation ponds

is shown in Figure 11-13. An operation and maintenance

cost curve is shown in Figure 11-14.

Table 11-10. Capital Cost Equation Determinants,

Evaporation Ponds (U.S. EPA, 1993c)

Brine Volume

(gpd)

Number of

Ponds

Total Surface

Area (acres)

 31,000 2 10

 50,000 2 15

100,000 2 30

200,000 4 60

400,000 6 120

500,000 3 150

Figure 11-14. O&M costs for evaporation ponds.

Figure 11-13. Capital costs for evaporation ponds.

195
Capital Cost

Y = [193,600 + 28.8(X)] + [4.08(X0.98)(Z)]

Y = [equipment cost] + [land cost]

where

Y=$

X = gallons of brine per day

Z = land cost in $/acre (e.g., $10,000/acre)

Range: 30,000 gpd ≤ X ≤ 150,000 gpd

Y = [44,820(X0.5) − (1.43 x 107

)] + [4.08(X0.98)(Z)]

Y = [equipment cost] + [land cost]

where

Y=$

X = gallons of brine per day

Z = land cost in $/acre (e.g., $10,000/acre)

Range: 150,000 gpd < X ≤ 500,000 gpd

Operation and Maintenance

Y = 11,600 + 0.96(X)

where

Y = $/year

X = gallons of brine per day

Range: 30,000 gpd ≤ X ≤ 150,000 gpd

Y = 1,500(X0.50) − 476,000

where

Y = $/year

X = gallons of brine per day

Range: 150,000 gpd < X ≤ 500,000 gpd

Total Annual Cost

Y = [CAP(CRF)] + O&M

where

Y = $/year

CAP = capital cost

CRF = capital recovery factor (e.g., 0.1175)

O&M = operation and maintenance cost

11.6 Discharge to Publicly Owned

Treatment Works

This section presents the design assumptions, capital

components, and the operation and maintenance components used to estimate the costs for discharging to a

POTW.

11.6.1 Design Assumptions

• Residuals flow rates range from 2,000 to 25 million

gallons per day, depending on the water treatment

technology.

• The sewer line connection is 500 feet or 1,000 feet

from the water treatment system.

• The residuals stream flows by gravity or under pressure from the treatment process to the sewer line.

• Two-inch minimum diameter pipe is used to prevent

clogging. Pipe diameter ranges from 2 to 24 inches.

• PVC or reinforced concrete piping is used.

• A 1- to 2-day storage lagoon is optional.

11.6.2 Capital Components

The capital components of system for discharge to a

POTW consist of piping and fittings, trenching, and land

clearing.

In addition, a pump, storage lagoon, electrical, and instrumentation are added for those facilities electing to

store residuals prior to discharge to a POTW.

Table 11-11 indicates the brine volumes, pipe diameters,

and lagoon capacities used to develop the capital cost

equation for discharge to POTW systems.

11.6.3 Operation and Maintenance

Components

The operation and maintenance components of each

system for discharge to a POTW include: labor and

supervision, maintenance labor and materials, basic

POTW charges, total suspended solids (TSS) surcharges, electricity, insurance, and general and administration.

11.6.4 Cost Components Excluded

The cost components not included in the cost equations

for systems for discharge to a POTW are any fees

charged by the POTW for the initial connection and land

cost.

It is assumed the water treatment facility would be able

to gain access for piping via an easement.

Table 11-11. Capital Cost Equation Determinants, Discharge

to POTW (U.S. EPA, 1993c)

Residuals Flow Rate

(gpd)

Pipe Diameter

(in.)

Lagoon

Capacity (gal)

 30,000 2 50,000

 67,500 2 115,000

 162,500 3 330,000

 500,000 4 750,000

 750,000 6 1,125,000

 1,000,000 6 1,500,000

10,000,000 24 15,000,000

25,000,000 36 37,500,000

196
11.6.5 Cost Equations and Cost Curves for

Discharge to Publicly Owned

Treatment Works

The cost equations for the capital components and operation and maintenance were developed by estimating

the costs for eight different brine flow rates. The capital

cost equation calculates the total capital cost (i.e., installed capital plus indirect capital costs). The total annual cost is calculated from the capital and operation

and maintenance costs obtained using the equations

presented below.

11.6.5.1 500 Feet of Discharge Pipe

Capital costs for 500 feet of discharge pipe are shown

in Figure 11-15. Operation and maintenance costs are

shown in Figure 11-16.

Capital Costs

Y = 4,500

where

Y=$

X = gallons of brine per day

Range: 2,000 gpd ≤ X < 150,000 gpd

Y = 4,600 + (1.9 x 10-3)(X)

where

Y=$

X = gallons of brine per day

Range: 150,000 gpd ≤ X ≤ 25,000,000 gpd

Operation and Maintenance

Y = 1,000 + [(1.83 x 10-3)(365)(X)] + [(0.12)(365)(Z)]

Y = [operating cost] + [POTW volume charge] +

[POTW strength charge]

where

Y = $/year

X = gallons of brine per day

Z = TSS pounds per day in excess of 300 ppm

(1.83 x 10-3 = POTW volume charge)

(0.12 = TSS surcharge)

Range: 2,000 gpd ≤ X ≤ 25,000,000 gpd

Total Annual Cost

Y = [CAP(CRF)] + O&M

where

Y = $/year

CAP = capital cost

CRF = capital recovery factor (e.g., 0.1175)

O&M = operation and maintenance cost

11.6.5.2 1,000 Feet of Discharge Pipe

Capital costs for 1,000 feet of discharge piping are

shown in Figure 11-17 and operation and maintenance

costs are shown in Figure 11-18.

Figure 11-15. Capital costs for 500 feet of discharge pipe. Figure 11-16. O&M costs for 500 feet of discharge pipe.

197
Capital Costs

Y = 8,700

where

Y=$

X = gallons of brine per day

Range: 2,000 gpd ≤ X < 150,000 gpd

Y = 9,000 + (3.9 x 10-3)(X)

where

Y=$

X = gallons of brine per day

Range: 150,000 gpd ≤ X ≤ 25,000,000 gpd

Operation and Maintenance

Y = 1,000 + [(1.83 x 10-3)(365)(X)] + [(0.12)(365)(Z)]

Y = [operating cost] + [POTW volume charge] +

[POTW strength charge]

where

Y = $/year

X = gallons of brine per day

Z = TSS pounds per day in excess of 300 ppm

(1.83 x 10-3 = POTW volume charge)

(0.12 = TSS surcharge)

Range: 2,000 gpd ≤ X ≤ 25,000,000 gpd

Total Annual Costs

Y = [CAP(CRF)] + O&M

where

Y = $/year

CAP = capital cost

CRF = capital recovery factor (e.g., 0.1175)

O&M = operation and maintenance cost

11.6.5.3 500 Feet of Discharge Pipe With a

Storage Lagoon

Capital and operation and maintenance cost curves for

500 feet of discharge pipe with a storage lagoon are

shown in Figures 11-19 and 11-20.

Capital Costs

Y = [44.3(X0.56)] + [0.5(Z)]

Y = [equipment cost] + [land cost]

where

Y=$

X = gallons of brine per day

Z = land cost in dollars per acre (e.g., $10,000/acre)

Range: 25,000 gpd ≤ X ≤ 750,000 gpd

Y = [1.13(X0.84)] + [(2.47 x 10-5)(X0.74)(Z)]

Y = [equipment cost] + [land cost]

where

Y=$

X = gallons of brine per day

Figure 11-17. Capital costs for 1,000 feet of discharge pipe. Figure 11-18. O&M costs for 1,000 feet of discharge pipe.

198
Z = land cost in dollars per acre (e.g., $10,000/acre)

Range: 750,000 gpd < X ≤ 25,000,000 gpd

Operation and Maintenance

Y = [3,500 + (2.25 x 10-2)(X)] + [(1.83 x 10-3)(365)(X)]

+ [(0.12)(365)(Z)]

Y = [operating cost] + [POTW volume charge] +

[POTW strength charge]

where

Y = $/year

X = gallons of brine per day

Z = TSS pounds per day in excess of 300 ppm

(1.83 x 10-3 = POTW volume charge)

(0.12 = TSS surcharge)

Range: 25,000 gpd ≤ X ≤ 500,000 gpd

Y = [0.30(X0.81)] + [(1.83 x 10-3)(365)(X)] +

[(0.12)(365)(Z)]

Y = [operating cost] + [POTW volume charge] +

[POTW strength charge]

where

Y = $/year

X = gallons of brine per day

Z = TSS pounds per day in excess of 300 ppm

(1.83 x 10-3 = POTW volume charge)

(0.12 = TSS surcharge)

Range: 500,000 gpd < X ≤ 25,000,000 gpd

Total Annual Cost

Y = [CAP(CRF)] + O&M

where

Y = $/year

CAP = capital cost

CRF = capital recovery factor (e.g., 0.1175)

O&M = operation and maintenance cost

11.6.5.4 1,000 Feet of Discharge Pipe With a

Storage Lagoon

Capital and operation and maintenance cost curves for

1,000 feet of discharge pipeline with a storage lagoon

are shown in Figures 11-21 and 11-22.

Capital Cost

Y = [19,500 + (0.11)(X)] + [0.5(Z)]

Y = [equipment cost] + [land cost]

where

Y=$

X = gallons of brine per day

Z = land cost in dollars per acre (e.g., $10,000/acre)

Range: 25,000 gpd ≤ X ≤ 750,000 gpd

Figure 11-20. O&M costs for 500 feet of discharge pipe with

storage lagoon.

Figure 11-19. Capital costs for 500 feet of discharge pipe with

storage lagoon.

199
Y = [60,200 + (6.97 x 10-2)(X)] + [(2.47 x 10-5)

(X0.74)(Z)]

Y = [equipment cost] + [land cost]

where

Y=$

X = gallons of brine per day

Z = land cost in dollars per acre (e.g., $10,000/acre)

Range: 750,000 gpd < X ≤ 25,000,000 gpd

Operation and Maintenance

Y = [3,500 + (2.25 x 10-2)(X)] + [(1.83 x 10-3)(365)(X)]

+ [(0.12)(365)(Z)]

Y = [operating cost] + [POTW volume charge] +

[POTW strength charge]

where

Y = $/year

X = gallons of brine per day

Z = TSS pounds per day in excess of 300 ppm

(1.83 x 10-3 = POTW volume charge)

(0.12 = TSS surcharge)

Range: 25,000 gpd ≤ X ≤ 500,000 gpd

Y = [0.30(X0.81)] + [(1.83 x 10-3)(365)(X)] +

[(0.12)(365)(Z)]

Y = [operating cost] + [POTW volume charge] +

[POTW strength charge]

where

Y=$/year

X=gallons of brine per day

Z=TSS pounds per day in excess of 300 ppm

(1.83 x 10-3 = POTW volume charge)

(0.12 = TSS surcharge)

Range: 500,000 gpd < X ≤ 25,000,000 gpd

Total Annual Cost

Y = [CAP(CRF)] + O&M

where

Y = $/year

CAP = capital cost

CRF = capital recovery factor (e.g., 0.1175)

O&M = operation and maintenance cost

11.7 Direct Discharge

This section presents the design assumptions, capital

components, and the operation and maintenance components used to estimate the costs for direct discharge

into surface water.

11.7.1 Design Assumptions

• Residuals flow rates range from 2,000 to 25,000,000

gallons per day depending on the water treatment

technology.

• The discharge point is 500 or 1,000 feet from the

water treatment system.

Figure 11-22. O&M costs for 1,000 feet of discharge pipe with

storage lagoon.

Figure 11-21. Capital costs for 1,000 feet of discharge pipe with

storage lagoon.

200
• The residuals stream flows by gravity or under pressure from the treatment process to the discharge

point.

• Two-inch minimum diameter pipe is used to prevent

clogging. Pipe diameter ranges from 2 to 24 inches.

• PVC or reinforced concrete piping is used.

• A 1- to 2-day storage lagoon is optional.

11.7.2 Capital Components

The capital components for each direct discharge system are piping and fittings, trenching, and land clearing.

In addition, a pump, storage lagoon, electrical, and instrumentation are added for those facilities electing to

store residuals prior to direct discharge.

Table 11-12 indicates the residuals flow rate, pipe diameters, and storage lagoon capacities used to develop

the capital cost equation for direct discharge.

11.7.3 Operation and Maintenance

Components

The operation and maintenance components for each

direct discharge system include: labor and supervision,

maintenance labor and materials, electricity, insurance,

and general and administration.

11.7.4 Cost Components Excluded

The cost components not included in the cost equations

for direct discharge systems are: NPDES permit application and monitoring costs, and land costs. It is assumed that the water treatment facility would be able to

gain access for piping via an easement.

11.7.5 Cost Equations for Direct Discharge

The cost equations for the capital components and operation and maintenance were developed by estimating

the costs for eight different residuals flow rates. The

capital cost equation calculates the total capital cost

(i.e., installed capital plus indirect capital costs). The

total annual cost is calculated from the capital and operation and maintenance costs obtained using the equations presented below.

11.7.5.1 500 Feet of Discharge Pipe

Capital Costs

Y = 4,500

where

Y=$

X = gallons of brine per day

Range: 2,000 gpd ≤ X < 150,000 gpd

Y = 4,600 + [(1.9 x 10-3)(X)]

where

Y=$

X = gallons of brine per day

Range: 150,000 gpd ≤ X ≤ 25,000,000 gpd

Operation and Maintenance

Y = 1,000

where

Y = $/year

X = gallons of brine per day

Range: 2,000 gpd ≤ X ≤ 25,000,000 gpd

Total Annual Cost

Y = [CAP(CRF)] + O&M

where

Y = $/year

CAP = capital cost

CRF = capital recovery factor (e.g., 0.1175)

O&M = operation and maintenance cost

11.7.5.2 1,000 Feet of Discharge Pipe

Capital Cost

Y = 8,700

where

Y=$

X = gallons of brine per day

Range: 2,000 gpd ≤ X < 150,000 gpd

Y = 9,000 + [(3.9 x 10-3)(X)]

where

Y=$

X = gallons of brine per day

Range: 150,000 gpd ≤ X ≤ 25,000,000 gpd

Table 11-12. Capital Cost Equation Determinants, Direct

Discharge (U.S. EPA, 1993c)

Residuals Flow

Rate (gpd)

Pipe Diameter

(in.)

Lagoon

Capacity (gal)

 30,000 2 50,000

 67,500 2 115,000

 162,500 3 330,000

 500,000 4 750,000

 750,000 6 1,125,000

 1,000,000 6 1,500,000

10,000,000 24 15,000,000

25,000,000 36 37,500,000

201
Operation and Maintenance

Y = 1,000

where

Y = $/year

X = gallons of brine per day

Range: 2,000 gpd ≤ X < 25,000,000 gpd

Total Annual Cost

Y = [CAP(CRF)] + O&M

where

Y = $/year

CAP = capital cost

CRF = capital recovery factor (e.g., 0.1175)

O&M = operation and maintenance cost

11.7.5.3 500 Feet of Discharge Pipe With a

Storage Lagoon

Capital Costs

Y = [44.3(X0.56)] + [0.5(Z)]

Y = [equipment cost] + [land cost]

where

Y=$

X = gallons of brine per day

Z = land cost in dollars per acre (e.g., $10,000/acre)

Range: 25,000 gpd ≤ X ≤ 750,000 gpd

Y = [1.13(X0.84)] + [(2.47 x 10-5)(X0.74)(Z)]

Y = [equipment cost] + [land cost]

where

Y=$

X = gallons of brine per day

Z = land cost in dollars per acre (e.g., $10,000/acre)

Range: 750,000 gpd < X ≤ 25,000,000 gpd

Operation and Maintenance

Y = 3,500 + [(2.25 x 10-2)(X)]

where

Y = $/year

X = gallons of brine per day

Range: 25,000 gpd ≤ X ≤ 500,000 gpd

Y = 0.30(X0.81)

where

Y = $/year

X = gallons of brine per day

Range: 500,000 gpd < X ≤ 25,000,000 gpd

Total Annual Cost

Y = [CAP(CRF)] + O&M

where

Y = $/year

CAP = capital cost

CRF = capital recovery factor (e.g., 0.1175)

O&M = operation and maintenance cost

11.7.5.4 1,000 Feet of Discharge Pipe With a

Storage Lagoon

Capital Cost

Y = [19,500 + (0.11)(X)] + [0.5(Z)]

Y = [equipment cost] + [land cost]

where

Y=$

X = gallons of brine per day

Z = land cost in dollars per acre (e.g., $10,000/acre)

Range: 25,000 gpd ≤ X ≤ 750,000 gpd

Y = [60,200 + (6.97 x 10-2)(X)] + [(2.47 x 10-5)

(X0.74)(Z)]

Y = [equipment cost] + [land cost]

where

Y=$

X = gallons of brine per day

Z = land cost in dollars per acre (e.g., $10,000/acre)

Range: 750,000 gpd < X ≤ 25,000,000 gpd

Operation and Maintenance

Y = 3,500 + [(2.25 x 10-2)(X)]

where

Y = $/year

X = gallons of brine per day

Range: 25,000 gpd ≤ X ≤ 500,000 gpd

Y = 0.30(X0.81)

where

Y = $/year

X = gallons of brine per day

Range: 500,000 gpd < X ≤ 25,000,000 gpd

Total Annual Cost

Y = [CAP(CRF)] + O&M

where

Y = $/year

CAP = capital cost

CRF = capital recovery factor (e.g., 0.1175)

O&M = operation and maintenance cost

202
11.8 Land Application

11.8.1 Liquid Sludge Land Application Cost

Components

This section presents the design assumptions, capital

components, and the operation and maintenance components used to estimate the costs for the land application of liquid residuals.

11.8.1.1 Design Assumptions

• The storage lagoon is constructed with a synthetic

membrane liner, overlaid by 12 inches of sand.

• The applied residuals consist of lime or alum residuals.

• The water treatment plant is 500 feet from the storage

lagoon; the residuals are gravity-fed to the lagoon.

• The lagoons are sized for 6-month storage.

• The application field is cultivated farmland. The

farmer uses a personally owned tractor and fuel to

till the land following the residuals application.

– Application by truck: The residuals are applied

twice a year, spring and fall, in a 1-inch layer.

– Application by sprinkler: The sprinkler system is a

portable system with aboveground piping and a

radius of 50 feet. The land application rate is 1/4

inch per day.

11.8.1.2 Capital Components

The capital components for land application consist of:

storage lagoon, piping, land, electrical, and sprinkler

system, when appropriate. Table 11-13 indicates the

residuals volume and pond surface areas used to develop the capital cost equation for the land application

systems.

11.8.1.3 Operation and Maintenance

Components

The operation and maintenance components for land

application include: electricity, labor and supervision,

sludge removal from storage pond, maintenance labor

and materials, insurance, general and administration,

and transportation and land application fees, when appropriate.

11.8.1.4 Cost Components Excluded

Costs for incorporation of the liquid residuals into the

soil, if necessary, are not included in the cost equations

for liquid residuals.

11.8.1.5 Liquid Residuals Land Application Cost

Equations and Cost Curves

The cost equations for the capital components and operation and maintenance were developed by estimating

the costs for six different residuals flow rates. The capital

cost equation calculates the total capital cost (i.e., installed capital plus indirect capital costs). Equipment

costs and land costs are separate components in the

capital cost equation to enable the user to exclude the

cost for purchasing land if it is not necessary. The total

annual cost is calculated based on the capital and operation and maintenance costs obtained using the equations presented below.

Sprinkler System

A capital cost curve for liquid residuals land application

using a sprinkler system is shown in Figure 11-23. Operation and maintenance costs are shown in Figure 11-24.

Capital Cost

Y = [120(X0.78)] + [(2.06 x 10-3)(X0.66)(Z)] + 

[(2.87 x 10-4)(X)(Z)]

Y = [equipment cost] + [pond land cost] + [application

field cost]

where

Y=$

X = gallons of residuals per day

Z = land cost in $/acres (e.g., $10,000/acre)

Range: 2,000 gpd ≤ X ≤ 500,000 gpd

Operation and Maintenance

Y = 81.7(X0.56)

where

Y = $/year

X = gallons of residuals per day

Range: 2,000 gpd ≤ X ≤ 60,000 gpd

Y = 8.63(X0.76)

where

Y = $/year

X = gallons of residuals per day

Range: 60,000 gpd < X ≤ 500,000 gpd

Table 11-13. Capital Cost Equation Determinants, Liquid

Residuals Land Application (U.S. EPA, 1993c)

Residuals Volume

(gpd)

Pond Surface Area

(acres)

 2,000 0.27

 10,000 0.92

 20,000 1.26

 60,000 2.64

100,000 3.72

500,000 10.97 

203
Total Annual Cost

Y = [CAP(CRF)] + O&M

where

Y = $/year

CAP = capital cost

CRF = capital recovery factor (e.g., 0.1175)

O&M = operation and maintenance cost

Trucking System

Capital and operation and maintenance cost curves for

liquid residuals land application with a residuals application truck are shown in Figures 11-25 and 11-26, respectively.

Capital Cost

Y = [104(X0.78)] + [(2.06 x 10-3)(X0.66)(Z)]

Y = [equipment cost] + [land cost]

where

Y=$

X = gallons of residuals per day

Z = land cost in $/acre (e.g., $10,000/acre)

Range: 2,000 gpd ≤ X ≤ 100,000 gpd

Operation and Maintenance

Y = 18.9(X0.96)

where

Y = $/year

X = gallons of residuals per day

Range: 2,000 gpd ≤ X ≤ 100,000 gpd

Total Annual Cost

Y = [CAP(CRF)] + O&M

where

Y = $/year

CAP = capital cost

CRF = capital recovery factor (e.g., 0.1175)

O&M = operation and maintenance cost

11.8.2 Dewatered Residuals Land

Application Cost Components

11.8.2.1 Design Assumptions

• The residuals are stockpiled on site.

• The residuals are transported off site for land application.

• A front-end loader is used to load the residuals for

transport. The front-end loader is owned by the WTP.

• The application field is within 25 miles of the plant.

• The application field is agricultural. The farmer growing crops on the field uses a personally owned tractor

and fuel to till the residuals into the soil.

• The application rate is 2 dry tons of residuals per acre.

Figure 11-23. Capital costs for liquid sludge land application. Figure 11-24. O&M costs for liquid sludge land application.

204
11.8.2.2 Capital Components

The capital components for dewatered residuals land

application consist of: land for residuals stockpile, and

land clearing. Table 11-14 indicates the residuals volume and storage pile surface area used to develop the

cost equations for the land application of dewatered

residuals.

11.8.2.3 Operation and Maintenance

Components

The operation and maintenance components for each

dewatered residuals land application include: labor and

supervision, residuals loading, and transportation and

land application fees.

11.8.2.4 Cost Components Excluded

Costs for incorporation of the dewatered residuals into

the soil, if necessary, are not included in the cost equations for dewatered residuals.

11.8.2.5 Dewatered Residuals Land Application

Cost Equations and Cost Curves

The cost equations for the capital components and operation and maintenance were developed by estimating

the costs for four different residuals flow rates. The

capital cost equation calculates the total capital cost

(i.e., installed capital plus indirect capital costs). Equipment costs and land costs are separate components in

the capital cost equation to enable the user to exclude

the cost for purchasing land if it is not necessary. The

total annual cost is calculated based on the capital

and operation and maintenance costs obtained using

the equations presented below. Capital and operation

and maintenance cost curves for dewatered residuals

land application are shown in Figures 11-27 and 11-28,

respectively.

Capital Cost

Y = [(111)(X0.50) − 4,700] + [0.5(Z)]

Y = [equipment cost] + [land cost]

where

Y=$

X = gallons of residuals per day

Z = land cost in $/acre (e.g., $10,000/acre)

Range: 2,000 gpd ≤ X < 2,500 gpd

Figure 11-25. Capital costs for trucking system. Figure 11-26. O&M costs for trucking system.

Table 11-14. Capital Cost Equation Determinants, Dewatered

Residuals Land Application (U.S. EPA, 1993c)

Residuals Volume

(gpd)

Storage Pile Surface

Area (acres)

 2,000 0.34

10,000 2.2 

20,000 4.5 

60,000 9.2 

205
Y = [(111)(X0.50) − 4,700] + {[(4.9 x 10-2)

(X0.50) − 2](Z)}

Y = [equipment cost] + [land cost]

where

Y=$

X = gallons of residuals per day

Z = land cost in $/acre (e.g., $10,000/acre)

Range: 2,500 gpd ≤ X < 60,000 gpd

Operation and Maintenance

Y = 800 + 28.2(X)

where

Y = $/year

X = gallons of residuals per day

Range: 2,000 gpd ≤ X ≤ 60,000 gpd

Total Annual Cost

Y = [CAP(CRF)] + O&M

where

Y = $/year

CAP = capital cost

CRF = capital recovery factor (e.g., 0.1175)

O&M = operation and maintenance cost

11.9 Nonhazardous Waste Landfill

11.9.1 Off-Site Nonhazardous Waste Landfill

Cost Components

This section presents the design assumptions used to

estimate the costs for offsite nonhazardous waste landfills.

11.9.1.1 Design Assumptions

• A commercial nonhazardous waste landfill is used for

residuals disposal.

• Transportation distance varies from 5 to 50 miles.

• There is no economy of scale for large waste volumes.

• All wastes pass the Paint Filter Liquids Test.

• The waste does not exhibit the characteristics of ignitability, corrosivity, reactivity, or toxicity.

11.9.1.2 Cost Components

The cost components for nonhazardous waste landfill

consist of commercial nonhazardous waste landfill tipping fee and transportation fees.

11.9.1.3 Offsite Nonhazardous Waste Landfill

Cost Equation and Cost Curve

The total operation and maintenance cost is based on

the disposal and transportation costs shown in the equation presented below. An operation and maintenance

cost curve for offsite nonhazardous waste landfill with

Figure 11-28. O&M costs for dewatered sludge land 

application.

Figure 11-27. Capital costs for dewatered sludge land 

application.

206
transportation distances of 5 miles and 50 miles is

shown in Figure 11-29.

Total Annual Cost

Y = [35(X)] + [(2.48 + 0.16Z)(X)]

Y = [disposal] + [transportation]

where

Y = $/year

X = tons of residuals requiring disposal/year

Z = transportation distance from 5 to 50 miles

11.9.2 Onsite Nonhazardous Waste Landfill

Cost Components

This section presents the design assumptions used to

estimate the costs for onsite nonhazardous waste landfills, the capital cost components, and the operation and

maintenance components.

11.9.2.1 Design Assumptions

• The landfill has a 20-year operating life.

• The landfill is a combination fill (i.e., a design combining below and above grade fills).

• The landfill containment system consists of 2 feet of

clay, a 30 mil HDPE liner, 1 foot of sand with a

leachate collection system, a geotextile filter fabric,

and 1 foot of native soil fill.

• The intermediate cover consists of slope and earthfill

soils.

• A ground-water monitoring system is installed and is

sampled biannually.

• The final cover consists of 1 foot of native soil fill, a

geotextile support fabric, a 30 mil PVC liner, 1 foot

of sand with drain tiles, a geotextile filter fabric, and

1.5 feet of topsoil.

• The postclosure care period is 30 years.

11.9.2.2 Capital Components

The capital components for each landfill consist of land,

land clearing, landfill excavation, composite (clay and

synthetic) liner, leachate collection system, groundwater monitoring system, equipment storage and maintenance buildings, visual screening berm, bulldozer,

truck, and inspection, testing and quality assurance.

11.9.2.3 Operation and Maintenance Components

The operation and maintenance components for each

landfill include labor and supervision, maintenance labor

and materials, intermediate cover, ground-water monitoring, leachate collection and treatment, and utilities,

including electricity and water.

11.9.2.4 Closure Components

The closure components for each landfill include final

cover system (including drain tiles), revegetation, and

inspection, testing, and quality assurance.

11.9.2.5 Postclosure Components

The postclosure components for each landfill include

ground-water monitoring, leachate collection and treatment, landscape maintenance, slope maintenance, and

annual inspection.

11.9.2.6 Cost Components Excluded

Siting and permitting costs are not included in the cost

estimates. Regulatory requirements for RCRA-regulated Subtitle D landfills vary from state to state. Permitting costs will vary based on the size of the landfill and

the local governing jurisdiction.

11.9.2.7 Onsite Nonhazardous Waste Landfill

Cost Equations and Cost Curves

The capital cost, operation and maintenance cost, closure cost, and postclosure cost equations for onsite

nonhazardous waste landfills are presented below. A

capital cost curve is shown in Figure 11-30. An operation

and maintenance cost curve is presented in Figure 11-

31. Figure 11-32 shows a closure cost curve, and Figure

11-33 an annual postclosure cost curve. Figure 11-29. O&M costs for off-site nonhazardous waste 

landfill.

207
Figure 11-30. Capital costs for onsite nonhazardous waste

landfill.

Figure 11-32. Closure costs for onsite nonhazardous waste

landfill.

Figure 11-31. O&M costs for onsite nonhazardous waste 

landfill.

Figure 11-33. Postclosure costs for onsite nonhazardous

waste landfill.

208
Capital Costs

Y = [8,079(X0.67)] + {[20.4 + 0.33(X0.50)](Z)}

Y = [landfill cost] + [land cost]

where

Y=$

X = tons of residuals requiring disposal per year

Z = land cost in $/acre (e.g., $10,000/acre)

Range: 2,500 TPY ≤ X ≤ 100,000 TPY

Operation and Maintenance

Y = [2,411(X0.50)] − 83,900

where

Y = $/year

X = tons of residuals requiring disposal per year

Range: 2,500 TPY ≤ X < 30,000 TPY

Y = [246(X0.50)] + 300,900

where

Y = $/year

X = tons of residuals requiring disposal per year

Range: 30,000 ≤ X ≤ 100,000 TPY

Closure

Y = 506(X0.80)

where

Y=$

X = tons of residuals requiring disposal per year

Range: 2,500 TPY ≤ X ≤ 100,000 TPY

Postclosure

Y = 600 + 58.2(X0.5)

where

Y = $/year

X = tons of residuals requiring disposal per year

Range: 2,500 TPY ≤ X ≤ 100,000 TPY

Total Annual Cost

Y = [8,483(X0.50) − 250,300] + {[2.4 + 0.04(X0.50)](Z)}

where

Y = $/year

X = tons of residuals requiring disposal per year

Z = land cost in $/acre (e.g., $10,000/acre)

Range: 2,500 TPY ≤ X ≤ 100,000 TPY

11.10 Hazardous Waste Landfill

This section presents the design assumptions, capital

components, and the operation and maintenance components used to estimate the costs for hazardous waste

landfills.

11.10.1 Design Assumptions

• A commercial hazardous waste landfill is used for

residuals disposal.

• Transportation distance varies from 200 to 500 miles.

• There is no economy of scale for large waste volumes.

• All wastes pass the Paint Filter Liquids Test or, if they

fail it, they are stabilized prior to disposal.

• The waste fails the TCLP test.

• Some wastes require stabilization prior to disposal to

meet the Land Disposal Restrictions Treatment

Standards for toxicity characteristic (TC) wastes.

11.10.2 Cost Components

The cost components for hazardous waste landfill consist of: hazardous waste landfill charges, stabilization

charges, and transportation fees.

11.10.3 Cost Components Excluded

The cost components not included in the cost equations

for hazardous waste landfills are: generator notification

requirements, manifest requirements, and other RCRA

requirements, as applicable.

11.10.4 Total Annual Cost Equation and Cost

Curves

Operation and maintenance cost equations for hazardous waste disposal and for stabilization and hazardous

waste disposal are presented below.

11.10.4.1 Hazardous Waste Disposal

An operation and maintenance cost curve for hazardous

waste disposal is shown in Figure 11-34, based on

transportation distances of 200 and 500 miles.

Y = [200 X] + {[7.9 + 0.22(Z)](X)}

Y = [disposal] + [transportation]

where

Y = $/year

X = tons of residuals requiring disposal per year

Z = transportation distance between 200 and 500

miles

11.10.4.2 Stabilization and Hazardous Waste

Disposal

An operation and maintenance cost curve for stabilization

and hazardous waste disposal is shown in Figure 11-35.

209
Y = [400(X)] + {[7.9 + 0.22(Z)](X)}

Y = [stabilization/disposal] + [transportation]

where

Y = $/year

X = tons of residuals requiring disposal per year

Z = transportation distance between 200 and 500

miles

11.11 Radioactive Waste Disposal

11.11.1 Low-Level Radioactive Waste

Disposal

There are three low-level radioactive waste disposal

sites operating in Nevada, South Carolina, and Washington, and one naturally occurring radioactive material

(NORM) disposal site operating in Utah. Each low-level

site bases waste disposal costs on several factors including state of origin of the waste, type of waste, and

measured radioactivity at the container surface. All disposal facilities are part of a regional pricing compact.

The Low-Level Radioactive Waste Policy Amendments

Act (LLRWPAA) of 1985 authorized the formation of

regional compacts and a system of incentives and penalties to ensure that states and compacts will be responsible for their own waste after January 1, 1993. States

that are part of a compact with an operational waste

disposal facility pay the lowest disposal costs while

states located outside of the compacts with disposal

facilities pay a disposal surcharge of $40 or $120 per

cubic foot (approximately $300 to $900 per 55-gallon

drum). The NORM disposal site is not part of a pricing

compact and does not add surcharges.

11.11.2 Cost Information for Radioactive

Waste Disposal

Radioactive waste disposal costs are not included within

this chapter. These costs may be obtained from EPA’s

guideline document, Suggested Guidelines for the Disposal of Drinking Water Treatment Wastes Containing

Naturally Occurring Radionuclides (U.S. EPA, 1990c).

These guidelines provide cost-related information, recommendations, and a summary of existing regulations

and criteria used by EPA and other agencies in addressing the disposal of radionuclides.

Figure 11-34. O&M costs for hazardous waste disposal.

Figure 11-35. O&M costs for stabilization and hazardous waste

disposal.

210
Chapter 12

Case Studies

12.1 Case Study 1: Disposal of Water

Treatment Residuals From Pine

Valley Water Treatment Plant,

Colorado Springs, Colorado

The Pine Valley Water Treatment Plant, located on the

southwest boundary of the U.S. Air Force Academy in

Colorado Springs, Colorado, was originally constructed

as a 42 million gallons per day (mgd) direct filtration

plant and placed into service in July 1969. The location

of the plant is shown in Figure 12-1. Expansion and

improvements over the years have increased design

and overload capacities to 56 and 84 mgd, respectively.

The addition of sedimentation basins in 1987 converted

the plant to a conventional treatment plant. Figure 12-2

illustrates the present site plan and facilities.

12.1.1 Facility Information

12.1.1.1 Treatment Processes

The Pine Valley Treatment Plant is a conventional facility

using the following processes: chemical addition with

in-line static mixing, coagulation-flocculation, sedimentation, and dual-media filtration, using two parallel treatment trains. An underground reservoir adjacent to the

plant provides 10 mgd of treated water storage. Two of

the eight filters are multimedia, i.e., containing coal, filter

sand, and garnet sand. The sedimentation basins contain stainless steel lamella plates. The rapid mix and

slow mix units, numbers 6 and 10 on Figure 12-3, have

been taken out of service. Figure 12-3 is a schematic

showing the direction of flow through the facility.

The chemicals applied at the Pine Valley Plant include

chlorine, aluminum sulfate, sodium aluminate, coagulant aid (Nalco 8100), filter aid (Nalco 8170), and soda

ash. Monthly average dosages for the chemicals used

from 1987 through 1992 are shown in Appendix D.

12.1.1.2 Raw Water Supply

The Pine Valley Plant was constructed primarily to treat

the Homestake Project water supply stored in Rampart

Reservoir. Subsequent acquisition of additional western

slope water supplies has provided other sources needing storage in the Rampart Reservoir. A 3.5-mile pressurized pipeline supplies raw water from the reservoir to

the plant. Water quality data for the plant’s influent raw

water for 1987 through 1992 are shown in Appendix D.

12.1.1.3 Plant Facility and Flow

Raw water is conveyed to the plant through a pipeline that

reduces from 42 to 36 to 30 inches (see Figure 12-2).

Modulating valves are in place to control the flow of water

to one of two parallel treatment trains designated as “A” or

“B.” Water enters a static mixer, continues through the

flocculation and sedimentation basins, and then flows onto

the filters, as shown in this schematic. Treated water leaving the plant is stored in an adjacent 10 million gallon (MG)

covered reservoir until needed in the system. Filter backwash water is discharged into the washwater recovery

basin. Sedimentation basin sludge and washdown water

are discharged to the sediment collection pond (see Figure

12-4) and/or washwater recovery basin. Water from the

recovery basin is systematically recycled to the plant headworks. Annual figures for the quantities of waste streams

generated are presented in Table 12-1. Decant water from

the sedimentation/collection pond will not be recycled until

a pump station is constructed; one is currently scheduled

for installation in 1996.

12.1.2 Residuals Management

The original plan for residuals management at the Pine

Valley WTP included the use of sediment lagoons (sediment drying beds), decanting, and landfilling. Decanting

operations were discontinued when the 1973 Colorado

Water Quality Control Act was promulgated. Pine Valley,

in service since 1969, did experience labor intensive

operations and limited capacity to store their annual

sediment production. In 1980, additional sediment handling facilities were installed to include sediment removal and transfer equipment in the washwater

recovery basin and two sediment drying beds. The new

facilities did not perform as planned, resulting in residuals containing too much water and incapable of being

removed with loaders and dump trucks (see Figure 12-5).

Because the facilities did not produce the desired results, a long-range plan was developed for ultimate

disposal. The plan, considered the most suitable and

211
economical method of sediment disposal, was to obtain

a site just east of Pine Valley and construct new disposal

facilities. The new sediment disposal facilities would

receive sediment from the Pine Valley WTP and another

proposed 150-mgd WTP. The new facilities, including a

175 acre-foot (acre-ft) sediment collection pond and two

3.5-acre-ft sediment drying ponds, were constructed

and placed into service in 1991. The new 150-mgd WTP

is scheduled for completion in 1996.

12.1.3 Residuals Handling Facilities and

Operations

12.1.3.1 Facilities

Current residuals disposal facilities at Pine Valley consist of a 1.5-MG washwater recovery basin with sediment collection and transfer equipment, four sediment

drying ponds (2 MG), and an offsite designated landfill

(Hanna Ranch). Additional facilities include sharing of

the 17 million-ft3

 sediment collection/holding basin, and

the two 3.5-acre-ft sediment drying ponds. Equipment

used to transfer sediment from the sediment drying

ponds to the landfill area include a sludge pump, tractor,

and 9,000-gallon tank trailer. Additional equipment is

used at the landfill site for monitoring, landscaping, and

erosion, dust, weed, and pest control.

12.1.3.2 Operations

Filter backwash water is discharged into the 1.5-MG

washwater recovery basin. Water containing sediment,

generated by washing down the flocculation and sedimentation basins, is transferred to the 128-MG sediment

collection basin (see Figures 12-2 and 12-4). Decant

from the recovery basin is pumped back to the plant

Figure 12-1. Locus map of Pine Valley WTP, Colorado Springs, CO (Colorado Springs Utilities, 1994).

212
Figure 12-2. Schematic of Pine Valley WTP, Colorado Springs, CO (Colorado Springs Utilities, 1994).

213
 1. Diversion dam (raw water intake)


 2. 980,000 Gallon Regulating Tank
 3. 36-Inch Raw Water Supply Line
 4. Plant Divides Into A (south) and B (north) Halves
 5. Control Valves
 6. Rapid Mixers
 7. Alum Feeders
 8. Soda Ash Feeders
 9. Chlorinators
10. Slow Mix Basins
11. Filter Influent Headers and Valves
12. Central Control Console

13. Diagrammatic Section Rapid Sand Filter
14. Cutaway View of Filter Bed
15. 42-Inch Filtered Water Header
16. 10 Million Gallon Filtered Water Reservoir (underground)
17. 48-Inch Transmission Main to City
18. Washwater Tank Supply
19. 400,000 Gallon Washwater Storage Tank
20. 24-Inch Washwater Waste Pipe
21. Washwater Recovery Basin and Pump Station
22. 20-Inch Recovery Basin Discharge Pipe 
 (returns wastewater to raw water system)
23. Sludge Drying Beds

Pump

Stream

ow Chart

To City

Pump

Filter Bottom

To Bed 
Drying 
Ponds 
Fig. C-1

Figure 12-3. Flow chart of Pine Valley WTP, Colorado Springs, CO (Colorado Springs Utilities, 1994).

214
headworks and retreated. The supernatant from the

sediment collection basin will be pumped either to the

recovery basin at Pine Valley or to the new WTP, when

operational. The sediment that is retained in the recovery basin is systematically pumped and sprayed onto

one of the four sediment drying ponds. Prior operations

and storage capacities required that the ponds be

cleaned annually. Present facilities will provide both

Pine Valley and the new WTP with sufficient capacity to

store their combined sediment production for a period of

15 to 20 years before removal is necessary.

12.1.4 Residuals Disposal

The first cleaning of the sediment drying ponds at Pine

Valley occurred in 1971. Data such as yardage, cost,

and disposal site for cleaning operations from 1971

through 1991 are presented in Table 12-2. (Note that in

1981, two more sediment drying ponds were added.)

Chemical characteristics of the residuals generated

from this process are represented in two water quality

laboratory reports from 1982 and 1992 (Tables 12-3 and

12-4). The 1982 analysis employed the Aqua Regea

total dissolving method, while the 1992 sample was

analyzed using the EP toxicity test.

From 1971 to 1976, cleaning and sediment disposal

operations were handled by the WTP’s construction and

equipment divisions. A dragline, trackloader, and loading

dump trucks were used to haul the frozen sediment to

an area on the U.S. Air Force Academy (USAFA)

grounds, 1/4 mile away. In 1977, the sediment drying

ponds were full well ahead of freezing and had to be

cleaned. The water construction division did not have

the equipment necessary to handle the soupy residue.

As a result, that was the year the department began

contracting sediment disposal operations. The contractor pumped the diluted sediment (2 to 4 percent solids)

into a tanker and hauled it to a private landfill, with the

exception of the first contractor who hauled it to the area

on the USAFA.

Sediment, accumulating at the rate of 7,000 to 8,000

cubic yards per year, eventually became too expensive

to contract. The Water Department purchased the

needed equipment (sludge pump and tanker) and, in

1983, began hauling sediment (5 to 10 percent solids)

itself. The sediment was hauled to Hanna Ranch (cityowned property), for which a certificate of designation

was obtained. With the new sediment disposal facilities

in operation and receiving most of the sediments, the

sediment drying ponds at Pine Valley did not require

cleaning in 1992.

12.1.5 Handling and Disposal Costs

Over the years, several thousands of dollars have been

spent in removing tons of sediment from Pine Valley.

Specifically, the costs associated with cleaning and

hauling sediment from Pine Valley have totaled almost

$363,000 since 1971 (see Table 12-2). The sludge pump

and tank trailer were purchased for $57,000. The 1981

additions to the disposal facilities cost over $475,000,

and the 1991 construction of the sediment disposal

facilities cost almost 5 million dollars (excluding engineering fees). Additional expenses and time were spent

obtaining certificates of designation for the Hanna

Ranch and Pine Valley disposal sites. Annual engineering

Figure 12-4. Sediment and decant piping schematic, Pine Valley WTP, Colorado Springs, CO (Colorado Springs

Utilities, 1994).

Table 12-1. Water Treated and Used, Pine Valley WTP,

Colorado Springs, CO (Pine Valley WTP, 1994)

Water

Treated

Washwater

Percent

 Washwater

Maintenance

Water

Percent

W/W &

M/W

Recycled 

Water

1987 9,805.310 165.185 1.71 12.420 1.83 361.0409

1988 11,636.362 192.959 1.66 18.687 1.82 224.1929

1989 11,820.405 142.346 1.20 26.609 1.43 227.6053

1990 10,112.920 115.852 1.13 17.552 1.32 211.0325

1991 10,082.701 117.096 1.16 18.890 1.35 234.9224

1992 11,010.706 131.890 1.20 15.995 1.34 261.4732

Averages 10,744.743 144.555 18.359 253.3778

Note: Figures in million gallons.

215
Figure 12-5. Letter regarding new sludge handling facility at Pine Valley WTP, Colorado Springs, CO (Colorado Springs Utilities,

1994).

216
Figure 12-5. Letter regarding new sludge handling facility at Pine Valley WTP, Colorado Springs, CO (Colorado Springs Utilities,

1994) (Continued).

217
Figure 12-5. Letter regarding new sludge handling facility at Pine Valley WTP, Colorado Springs, CO (Colorado Springs Utilities,

1994) (Continued).

218
costs for certification renewal since 1991 are unavailable. Additional operation and maintenance charges not

accounted for include pest and weed control, landscaping, erosion control measures, and monitoring and collection data requirements.

12.2 Case Study 2: Disposal of Water

Treatment Residuals From Mesa

Treatment Plant, Colorado Springs,

Colorado

The Mesa Water Treatment Plant, located near the Garden of the Gods in the western part of Colorado Springs,

was built in 1942 with a design capacity of 4 mgd. It

treats waters from the Pikes Peak North and South

Slope watersheds. Several expansions and modifications since the plant’s opening have increased its design

and overload capacities to 42 and 64 mgd, respectively.

Figure 12-6 shows the site plan and facilities.

12.2.1 Facility Information

12.2.1.1 Treatment Processes

The existing treatment plant is a conventional facility

using chemical addition with rapid mixing, coagulation,

flocculation, and sedimentation in three separate treatment trains. Filtration of the combined discharge from

the three trains occurs in the filter building using eight

dual-media filters (see Figure 12-6). Chemical additions

Table 12-2. Sediment Disposal, Pine Valley WTP, Colorado Springs, CO (Pine Valley WTP, 1994)

Year Dates Yardage Contractor Cost Disposal Site

1969 N/A

1970 N/A

1971 02/17 2,500–3,000 Water Const. $3,600.00 USAFA

1972 01/17 2,500–3,000 Water Const. $3,750.00 USAFA

1973 01/05 2,500–3,000 Water Const. $3,900.00 USAFA

1974 02/26 2,500–3,000 Water Const. $4,200.00 USAFA

1975 02/28 2,500–3,000 Water Const. $4,350.00 USAFA

1976 04/20, 12/07 2,500–3,000 Water Const. $4,500.00 USAFA

1977 Nov. 2,057 Waste Transport $4,155.00 USAFA

1978 Aug. 3,491 JD Dye Landfill Costs $17,105.90 

$2,820.80

Landfill

1978 Sept. 3,609 JD Dye Landfill Costs $13,630.28 

$2,915.57

Landfill

1979 Apr. 2,997 JD Dye $14,863.50 Mesa

1979 Sept. 3,961 Ready Mix $19,370.17 Private

1980 May 3,900 Ready Mix $17,511.00 Private

1980 Aug. 4,900 Ready Mix $19,355.00 Private

1981 N/A (added two new beds)

1982 Mar. 7,100 JD Dye $41,110.15 Mesa

1983 Apr. 7,200 Water Const. $35,000.00 Hanna

1984 Sept. 7,500 Water Const. $33,000.00 Hanna

1985 Oct. 7,100 Water Const. $27,000.00 Hanna

1986 Sept. 7,100 Water Const. $28,400.00 Hanna

1987 Apr. 7,100 Water Const. $28,500.00 Hanna

1988 Mar. 2,600 Water Const. $10,400.00 Hanna

1988 July 1,800 Water Const. $7,200.00 Hanna

1989 N/A

1990 N/A

1991 Jan. 2,400 Water Const. $10,800.00 Hanna

1991 May 1,200 Water Const. $5,400.00 Mesa

1992 N/A

219
at the Mesa Plant include chlorine, aluminum sulfate,

sodium aluminate, coagulant aid (Nalco 8100), filter aid

(Nalco 8170), and soda ash. Monthly average dosages

for the chemicals used from 1987 through 1992 can be

found in Appendix E.

12.2.1.2 Raw Water Supply

Presently, the plant treats a combination of waters from the

Pikes Peak watershed, as well as from western slope

waters diverted into the North Slope Reservoirs. Water is

conveyed to the plant through three pipelines—two gravity

flow pipelines from the Manitou Hydro Afterbay, and one

pumped pipeline from the Fountain Creek intake pumping station. Water quality data for the plant influent raw

water from 1987 through 1992 are shown in Appendix

E. Results of influent laboratory analyses for April and

August of 1992 are listed in Tables 12-5 and 12-6. Data

from specific pipeline supplies are outlined in Table 12-7.

12.2.1.3 Plant Facility and Flow

Raw water enters the plant through three pipelines (see

Figure 12-6). Modulating valves control the flow of water

to one of three parallel treatment trains designated as

numbers 24, 30A, and 30B. Water enters one of the

rapid mix chambers, then travels through the flocculation and sedimentation basins before moving on to the

filters. Treated water leaving the plant is stored in an

adjacent 20-MG covered reservoir. Filter backwash

water and other wastewater from the basins are discharged into an 18-MG earthen collection basin and are

then recycled to the plant influent for retreatment. The

volumes of water recycled from 1987 through 1992 are

shown in Table 12-8.

12.2.2 Residuals Management

The management of water treatment plant residuals

commenced with the Federal Water Pollution Control

Act/Colorado Water Quality Control Act promulgated in

1973. This act prohibited the discharge of pollutants in

excess of prescribed limits into streams. Direct discharges of residuals from the Mesa Plant did not meet

the established standards; therefore, the practice was

discontinued. Residuals management and disposal was

then a reality and had to be dealt with.

Over the years, several contingency plans were studied

to determine the most effective and efficient method of

residuals management. This meant considering ways to

minimize operating expenses, reduce handling costs,

and maximize water reuse. Sediment lagoons (sediment

drying beds) were determined to be the most effective

way to handle residuals. Until this option could be implemented, certificates of designation were obtained or

amended, and alternate disposal sites were sought.

Other proposals that were occasionally considered included various types of mechanical dewatering equipment, chemical conditioning of the sediment, and

discharging wastes to a sanitary sewer. None of these

proposals ever materialized.

12.2.3 Residuals Handling Facilities and

Operations

12.2.3.1 Facilities

Current residuals disposal facilities at Mesa Plant consist of a large collection and holding basin (18 MG-washwater recovery basin), a sediment drying pond (4 MG),

Table 12-3. Characteristics of Residuals Generated in 1982,

Pine Valley WTP, Colorado Springs, CO (Pine

Valley WTP, 1994)

Water Utility: Colorado Springs Water Dept.

County: El Paso

Source of Supply: Pine Valley Sludge

Where Collected: Sludge Ponds

Collected By: John Uhrich

Date Sampled: 9/16/82

Date Analyzed: 9/20/82

Chemist: Monte Fryt

MAIL TO:

% Dry Wt = 1.07

Resultsa

Parameters

% Dry

Weight mg/L mg/g

Arsenic 0.0003 0.064 2.78

Barium 0.052 10.4 452.19

Cadmium 0.0007 0.14 6.1

Chromium 0.003 0.5 21.7

Fluoride 0.15 30.5 1,326

Lead 0.007 1.50 65.2

Mercury 0.00009 0.0175 0.76

Nitrate as N 0.09b N/A

Total Hardness as

CaC03

60b N/A

Sulfate 22.3b N/A

pH 6.62b N/A

Selenium 0.00003 0.006 0.26

Silver 0.0000003 0.00054 0.023

Calcium as CaC03 0.65 130.0 5,652

Magnesium 0.075 15.0 652.2

Iron 2.01 400.0 17,391

Manganese 1.81 360.0 15,652

Zinc 0.66 132.0 5,739

Molybdenum 0.0007 0.158 6.87

PCB <0.0025 <0.01 <0.5

Aluminum 0.078 15.5 674

a Factor converting mg/L to μg/g = 43.48. b As supernatant.

220
Table 12-4. Characteristics of Residuals Generated in 1992, Pine Valley WTP, Colorado Springs, CO (Pine Valley WTP, 1994)

CITY OF COLORADO SPRINGS WATER QUALITY LABORATORY

SAMPLE REPORT

Sample Site: PVSD Pine Valley Sediment (PVSD) Date: 3/18/93 09:56 AM

Sample Number: 92120331 Water Type: M 

Comments: Sample Date: 12/30/92

Analysis Description/Code Result Result Code

Minimum

Reported

Level Test Units Method Code

pH PH 5.8 S.U. E

Percent solid PS 14.8 mg/L GRAV

Magnesium (dissolved) MGD 0.16 0.10 mg/L FAASA

Silver (dissolved) AGD 0 ND 0.5 μg/L GFAAS

Aluminum (dissolved) ALD 3,200 400 μg/L FAASN

Arsenic (dissolved) ASD 0 ND 4 μg/L GFAAS

Barium (dissolved) BAD 150 100 μg/L FAASN

Cadmium (dissolved) CDD 2.2 0.10 μg/L GFAAS

Chromium (dissolved) CRD 0 ND 50 μg/L FAASN

Copper (dissolved) CUD 19 12 μg/L FAASA

Mercury (dissolved) HGD 0 ND 0.2 μg/L CVAAS

Lead (dissolved) PBD 2 1.0 μg/L GFAAS

Selenium (dissolved) SED 0 ND 1.0 μg/L VGAAS

Key

ND = Not detected.

Operation


Control


Center

Alum


Coag. Aid


Sod. Alum.

Rapid Mix

Rapid Mix

Settling Basin


(30B)

Settling Basin


Flocculation


Basin


(30A)

Alum


Coag. Aid


Sod. Alum.

No Scale

Washwater


Recovery basin Drain Line

Solids


Transfer 


Line

Drain Line

Washwater 


Drain Line

Washwater


Pump Station


2 - Total Cap.


4:0 MGD

Wash-


water 


Tank

Filter Aid

30"

30"

Filter 


Effluent

Chlorine


Soda Ash

Soda Ash


Storage and Feed 


Facilities (Dry System)

Highline

Reservoir


20 MG

Clarifier


Basin


(24)

Control


Room

Filters

Filter Building

Filter


Effluent

Received Washwater

Flow Tube

Access Road

Chemical 


Building

Flow 


Meter

Access Road

Chlorine

12" Mesa Road

30" From 33rd St


30" From Hydro


20" From Hydro

16"

24"

42"

To Sed


Drying


Bed

Flocculation


Basin

Flocculation


Basin

Rapid Mix

Figure 12-6. Mesa WTP, Colorado Springs, CO (Colorado Springs Utilities, 1994).

221
and an onsite designated landfill area. Equipment used

to transfer sediment from the holding basin to the landfill

area includes a Mudcat/dredge, trackloader, dump

trucks, and bulldozer.

12.2.3.2 Operations

Spent filter backwash water and water latent with sediment from the flocculation and sedimentation basins are

discharged into the recovery basin. Decant water is

periodically pumped back to the head of the plant and

retreated. The sediment that is retained in the wastewater recovery basin accumulates at the rate of approximately 7,000 to 8,000 cubic yards per year and remains

there until removed. Mesa has sufficient capacity to

store approximately 40,000 to 50,000 cubic yards of

sediment in its recovery basin before storage becomes

a problem. This storage capacity is equivalent to roughly

4 to 5 years of residuals production. Analyses of the

chemical characteristics of the residuals generated in

1978 and 1992 are presented in Tables 12-9A, 12-9B,

and 12-10, respectively. Note that there are two different

Table 12-5. Laboratory Samples: April 1992, Fountain Creek, Colorado Springs, CO (Mesa WTP, 1994)

CITY OF COLORADO SPRINGS WATER QUALITY LABORATORY

SAMPLE REPORT

Sample Site: FCR Fountain Creek at 33rd Street (FCR) Date: 31-Mar-1993 03:59 PM

Sample Number: 92040281 Water Type: R

Comments: Sample Date: 28-APR-92

Analysis Description/Code Result Result Code

Minimum

Reported

Level Test Units Method Code

pH PH 8.07 S.U. E

Conductivity CON 245 μmhos E

Turbidity TUR 23 NTU N

HCO3 alkalinity ALK 81 mg/L T

CO3 alkalinity CO3 0 mg/L T

Hardness as CaCO3 HAR 94 mg/L T

Fluoride F 2.56 0.10 mg/L E

Total dissolved solids TDS 161 mg/L GRAV

Ammonia (total) NH3 0.06 0.05 mg/L E

Chloride CL 13 0.4 mg/L IC

Nitrate NO3 0.66 0.01 mg/L IC

Sulfate SO4 13.2 1.0 mg/L IC

Barium (dissolved) BAD 67 5 μg/L GFAAS

Cadmium (dissolved) CDD 0.26 0.10 μg/L GFAAS

Copper (total) CU 1 1 μg/L GFAAS

Iron (total) FE 980 40 μg/L FAASA

Iron (dissolved) FED 120 40 μg/L FAASA

Manganese (total) MN 120 20 μg/L FAASA

Manganese (dissolved) MND 30 20 μg/L FAASA

Selenium (dissolved) SED 0 ND 1.0 μg/L VGAAS

Chlorine residual CL2 0 mg/L DPD

Temperature centigrade TEC 8 C THER

Threshold odor number TON 3 NO

Total coliform TC 24,000 cfu/100mL MTF

Fecal coliform FC 5,000 cfu/100mL MTF

Fecal Streptococus FS 70 cfu/100mL MF

Key

ND = Not detected.

222
test methods employed in these analyses—the Aqua

Regea total dissolving method used in 1978, and the EP

toxicity test used in 1992.

12.2.4 Residuals Disposal

Since the first cleaning of the recovery basin in 1974,

the basin has been cleaned every 5 years. Information

pertaining to these cleaning operations can be found in

Table 12-11. In late 1973, the plant was taken out of

service, recycling most of the water from the recovery

basin before shutting down. Over the winter, the recovery basin was allowed to dry and freeze; in early spring,

a track loader, dragline, and dump trucks removed and

transferred the sediment to a city-owned landfill area.

In 1979, a different method of sediment disposal was

employed. Higher levels of water demand prevented the

Table 12-6. Laboratory Samples: August 1992, Mesa WTP, Colorado Springs, CO (Mesa WTP, 1994)

CITY OF COLORADO SPRINGS WATER QUALITY LABORATORY

SAMPLE REPORT

Sample Site: MPI Mesa Plant Influent (MPI) Date: 31-Mar-1993 04:01 PM

Sample Number: 92080222 Water Type: R

Comments: recov. pumps on; plant TUR=3.0 Sample Date: 25-AUG-92

Analysis Description/Code Result Result Code

Minimum

Reported

Level Test Units Method Code

pH PH 7.45 S.U. E

Conductivity CON 50 μmhos E

Color COL 25 COMP

Turbidity TUR 3.8 NTU N

HCO3 alkalinity ALK 16 mg/L T

CO3 alkalinity CO3 0 mg/L T

Hardness as CaCO3 HAR 24 mg/L T

Fluoride F 2.28 0.10 mg/L E

Chloride CL 0.4 0.4 mg/L IC

Nitrate NO3 0.13 0.01 mg/L IC

Sulfate SO4 3.3 1.0 mg/L IC

Sodium (dissolved) NAD 2.1 0.5 mg/L IC

Potassium (dissolved) KD 1.3 0.5 mg/L IC

Magnesium (dissolved) MGD 0.7 0.5 mg/L IC

Calcium (dissolved) CAD 5.8 3 mg/L IC

Silver (dissolved) AGD 0 ND 0.5 μg/L GFAAS

Arsenic (dissolved) ASD 0 ND 4 μg/L GFAAS

Chromium (dissolved) CRD 0 ND 0.4 μg/L GFAAS

Iron (total) FE 240 40 μg/L FAASA

Iron (dissolved) FED 120 40 μg/L FAASA

Manganese (total) MN 20 20 μg/L FAASA

Manganese (dissolved) MND 0 ND 20 μg/L FAASA

Chlorine residual CL2 0 mg/L DPD

Temperature centigrade TEC 12 C THER

Threshold odor number TON 1 NO

Total coliform TC 170 Cfu/100mL MF

Fecal coliform FC 4 Cfu/100mL MF

Fecal Streptococus FS 20 Cfu/100mL MF

Key

ND = Not detected.

223
plant and recovery basin from being taken out of service.

A floating dredge (Mudcat) was purchased to dredge and

pump the sediment to a small berm area isolated from one

end of the recovery basin. From here, a contractor pumped

the diluted sediment, which was mostly water, into a tanker

and hauled it to a private landfill.

Cleaning and sediment disposal operations in 1984 and

1989 were handled in the same manner—the Mudcat

was used to dredge and pump sediment from the recovery basin into a large sediment drying pond built in 1984.

A floating outlet discharge line was used to decant clear

water from the drying pond back into the recovery basin.

Cleaning operations were discontinued when the recovery basin was considered cleaned or when the drying

pond was full of sediment. The Mudcat was removed

from the recovery basin, and the sediment drying pond

was allowed to freeze, thaw, and dry over the winter and

into the next year.

In the spring of 1985, the underlying sediment was still

too soupy to load and haul. In autumn of 1985 and 1990,

after 1 year of the freeze-thaw drying cycle, the sediment had dewatered to a level conducive to hauling

(estimated between 10 to 15 percent solids). The depth

of sediment had been reduced from roughly 8 feet to 3

feet. A track loader, belly dump, and bulldozer were used

to transfer and spread the sediment onto the adjacent

landfill in a layer approximately 1 foot thick.

When the top layer of sediment at the landfill dried, it

would turn into a fine granular powder, which created

quite a dust problem when the wind blew. A dust ban

polymer was applied to minimize the dust problem during the dewatering process. In addition, newly placed

sediment at the landfill was seeded with native grasses

to control erosion and dust pollution. Ground-water

monitoring wells are not required at Mesa; however, they

are in place at Hanna Ranch where sediment from Mesa

has been taken. Results of a leachate sample collected

in April 1990 are shown in Figure 12-7.

12.3 Case Study 3: Land Application of

Water Treatment Plant Residuals at

Cobb County-Marietta Water

Authority, Marietta, Georgia

The Cobb County-Marietta Water Authority (CCMWA)

operates two WTPs in Cobb County, Georgia, which is

in the metropolitan Atlanta area. The James E. Quarles

Plant is permitted to treat 58 mgd of raw water from the

Chattahoochee River. The Chattahoochee River is the

primary source of drinking water for metropolitan Atlanta.

The Hugh A. Wyckoff Plant is permitted to treat 72 mgd

of raw water from Lake Allatoona, a U.S. Army Corps of

Engineers impoundment on the Etowah River in northwest Georgia.

Both plants use the same method of residuals handling.

Sludge is removed daily from the settling basins and

sent to gravity sludge thickeners. Backwash water from

the filters is sent to a backwash recycle tank where the

sludge settles out and the clarified backwash water is

returned to the head of the plant. The settled sludge is

scraped to one end of the recycle tank and pumped to

the gravity sludge thickeners. The gravity thickened

sludge is then pumped to a conditioning tank where lime

slurry is mixed with the sludge on a 10 to 15 percent dry

weight basis. The conditioned sludge is then dewatered

in a pressure filter to obtain a sludge cake of approximately 35 percent total solids. The cake is then transported to a sludge storage yard where it is stockpiled for

several months. Two to three times each year, the

sludge is loaded into trucks and hauled to agricultural

land in or near Cobb County. The dewatered sludge is

applied to the fields with a manure spreader. The

Table 12-7. Raw Water Source for Mesa WTP,

Colorado Springs, CO (Mesa WTP, 1994)

Fountain Creek

Sample

Manitou Afterbay

Sample

Mean Range Mean Range

pH 7.58 8.15–7.19 7.56 8.05–7.17

Alkalinity (mg/L) 27.30 76.0–11.0 24.5 62.0–12.0

Hardness (mg/L) 29.8 86.0–13.0 26.4 64.0–14.0

Temperature (°F) 47.2 62.7–35.1 47.0 62.7–35.6

Turbidity (NTU) 4.70 133.6–0.6 3.70 36.0–0.6 

Table 12-8. Water Treated and Used, Mesa WTP, Colorado Springs, CO, in Million Gallons (Mesa WTP, 1994)

Water

Treated Washwater

Percent

Washwater

Maintenance

Water

Percent

W/W & M/W

Recycled

Water

1987 6,089.827 149.977 2.46 2.729 2.51 319.408

1988 6,695.425 196.539 2.94 6.830 3.04 431.176

1989 6,554.006 269.520 4.11 8.359 4.24 388.040

1990 6,662.514 215.598 3.24 11.069 3.40 344.745

1991 6,929.688 201.402 2.91 19.709 3.19 389.142

1992 6,530.272 132.187 2.02 7.587 2.14 366.373

Averages 6,576.955 194.202 9.381 373.147

224
application rate depends on the lime requirements of the

soil and the lime content of the sludge.

As of 1995, there are no state or local regulations that

deal with the land application of WTP residuals. The

Georgia Environmental Protection Division has classified the residuals as “recovered materials” under the

Rules for Solid Waste Management, Chapter 391.3 to

4.04(7). Under this classification the residuals do not

require a solids waste handling permit.

12.3.1 Facility Information

The James E. Quarles Plant is presently permitted for a

total flow of 58 mgd. It consists of an older plant, Plant

No. 1, that is permitted to handle 36 mgd, and a newer

plant, Plant No. 2, that is permitted to handle 22 mgd.

Figure 12-8 is a schematic outline of the plant.

The raw water for both plants is drawn from the Chattahoochee River and pumped 8,500 feet to a small raw

water reservoir with a capacity of 20 MG. From the

reservoir, the water flows through separate lines to

Plants No. 1 and 2.

Plant No. 1 adds alum, chlorine, and chlorine dioxide at

the rapid mix facility. The water then flows through eight

hydraulic flocculators with a detention time of 17 minutes at the rated flow of 36 mgd. There are eight rectangular sedimentation basins with a theoretical detention

Table 12-9A. Chemical Characteristics of Residuals (Sludge) From Mesa WTP, Colorado Springs, CO, 1978 (Mesa WTP, 1994)

Water Analysisa

Location: Mesa Plant Sludge (6.571% dry wt) Date: 20 March, 1978

Analysis (Dry Wt) Limits Results Analysis (Dry Wt) Limits Results

Total coliform (per 100 mL) 4 — Iron (Fe) 1.30 0.30 1,600

Carbon dioxide (CO2) NL — Manganese (Mn) 0.22 0.05 270.0

Residual Cl2 NLT

 0.1

— Nickel (Ni) 0.001 NL 1.60

Temperature (°F) 78° — Copper (Cu) 0.012 1.0 14.6

Color intensity 15 — Lead (Pb) 0.000002 0.05 0.003

Taste and odor (Threshold

Odor Units)

 3 — Zinc (Zn) 0.014 5 17.0

Turbidity (NTU) 1.0 — Calcium (Ca) 0.23 NL 280.0

Specific conductance

(Micromhos –25°C)

NL — Magnesium (Mg) 0.12 125 145

Total dissolved solids NL — Cadmium (Cd) 0.0002 0.01 0.20

Hardness as CaCO3 NL — Barium (Ba) 0.003 1.0 3.50

Noncarbonate hardness as

CaCO3

NL — Sodium (Na) 1.82 NL 2,250

Bicarbonate (HCO3) NL — Potassium (K) 1.30 NL 1,600

Carbonate (CO3) NL — pH 9.5 

 5.6 

—

Chloride (Cl) 250 — Acidity NL —

Silica (SiO2) 0.085 NL 104.7 Mercury (Hg) 0.002 (μg/100 mL) 0.2 3.0

Sulfate (SO4) 0.28 250 350 Molybdenum (Mo) —

Nitrate (NO3) 45 — Arsenic (As) 0.0001 0.05 0.14

Total phosphate (PO4) 0.17 NL 205.4 Cyanide (CN) 0.0001 0.2 0.14

Fluoride (F) 0.22 62°

 2.0

 272 Fecal coliform (per 100 ml) 4 —

Aluminum (Al) 3.08 NL 3,800 Fecal Streptococcus (per 100 mL) —

Total chromium (Cr) 0.006 0.05 7.8 Plate count (org/mL)b Backwash

H2O

700

Sludge

pond

39

Cobalt (Co) 0.0004 0.50

a All chemical analyses in milligrams per liter except where otherwise indicated.

b Sample collected 3/22/78.

225
time of 160 minutes. The sludge is removed with retrofitted siphon-type sludge collectors located in the upper

one-third of each basin. Once or twice each year the

basins are drained to remove any buildup of sludge that

occurs.

The water is then filtered through eight double-bay gravity dual media filters. The present permitted filter rate is

4.5 gallons per minute per square foot (gpm/ft2

). The

filter backwash system consists of surface sweeps and

underdrain backwash.

The filtered water is then treated with lime, fluoride, and

additional chlorine before storage in the plant clearwells.

The finished water is then pumped to customers.

Plant No. 2, which was built in 1980, has the same

process train as Plant No. 1. Alum, chlorine, and chlorine dioxide are added at the rapid mix facility. The

water then flows through four hydraulic flocculators

with a detention time of 19 minutes at the rated flow

of 22 mgd. There are four rectangular sedimentation

basins with a theoretical detention time of 155

Table 12-9B. Chemical Characteristics of Residuals (Supernatant) From Mesa WTP, Colorado Springs, CO, 1978 (Mesa WTP, 1994)

Water Analysisa

Location: Mesa Plant Sludge (Supernatant) (93.429%) Date: 20 March, 1978

Analysis (Dry Wt) Limits Results Analysis (Dry Wt) Limits Results

Total coliform (per 100 mL) 4 Backwash

H2O

Sludge pond

20,000

Iron (Fe) 0.30 1.5

Carbon dioxide (CO2) NL 2.7 Manganese (Mn) 0.05 20.0

Residual Cl2 NLT

0.1

0 Nickel (Ni) NL 0.014

Temperature (°F) 78° — Copper (Cu) 1.0 0.009

Color intensity 15 — Lead (Pb) 0.05 0.001

Taste and odor (Threshold 

Odor Units)

 3 — Zinc (Zn) 5 0.025

Turbidity (NTU) 1.0 — Calcium (Ca) NL 41.6

Specific conductance

(Micromhos –25°C)

NL 260 Magnesium (Mg) 125 3.5

Total dissolved solids NL 174 Cadmium (Cd) 0.01 0.001

Hardness as CaCO3 NL 109 Barium (Ba) 1.0 0.002

Noncarbonate hardness as

CaCO3

NL 40 Sodium (Na) NL 11.2

Bicarbonate (HCO3) NL 69 Potassium (K) NL 4.0

Carbonate (CO3) NL 0 pH 9.5

5.6

7.70

Chloride (Cl) 250 8.5 Acidity NL 3.5

Silica (SiO2) 0.085 NL 1.55 Mercury (Hg) (μg/100 mL) 0.2 0.062

Sulfate (SO4) 0.28 250 30 Molybdenum (Mo) —

Nitrate (NO3) 45 0.25 Arsenic (As) 0.05 0.004

Total phosphate (PO4) NL 0.55 Cyanide (CN) 0.2 0

Fluoride (F) 62°

2.0

2.40 Fecal coliform (per 100 mL) 4 —

Aluminum (Al) NL 0.17 Fecal Streptococcus (per 100 mL) —

Total chromium (Cr) 0.05 0.01 Plate count (org/mL)b Backwash

H2O

700

Sludge 

pond

39

Cobalt (Co) 0.0025

a All chemical analyses in milligrams per liter except where otherwise indicated.

b Sample collected 3/22/78

226
minutes. The sludge is removed with siphon-type

sludge collectors that traverse the entire length of

each basin. The sedimentation basins are drained

once or twice each year to remove any sludge buildup.

The water is then filtered through four double-bay gravity

dual media filters. The present permitted filter rate is 5.5

gpm/ft2

. The filter backwash system consists of surface

sweeps and underdrain backwash. One filter is fitted

with an air backwash system.

The filtered water is then treated with lime, fluoride, and

additional chlorine before storage in the plant clearwells.

The finished water is then pumped to customers.

The Hugh A. Wyckoff Plant is presently permitted for a

total flow of 72 mgd. The raw water for this plant is drawn

from Lake Allatoona and pumped 4.5 miles to the plant.

Figure 12-9 is a schematic outline of the plant.

Alum, chlorine, and chlorine dioxide are added before

an in-line mechanical mixer. The water then goes

through six flocculator basins. Four basins have mechanical mixers with a detention time of 23 minutes. Two

basins have hydraulic mixing with a detention time of 32

minutes. Six rectangular sedimentation basins have a

detention time of 136 minutes. The sludge is removed

from four of the basins with circular mechanical sludge

collectors, located in the first two-thirds of each basin.

The two newer basins are equipped with floating siphontype sludge collectors for the entire length of the basins.

The sedimentation basins are drained each year to remove sludge buildup.

Table 12-11. Sediment Disposal, Mesa WTP, Colorado Springs, CO (Mesa WTP, 1994)

Year Date Cubic Yards Contractor Cost Disposal Site Method

1974 Mar. 40,000 Water construction $17,000 City landfill Loader and dump trucks

1979 May 50,000 Dye construction $150,000 Private landfill Dredge and tanker

1984 Oct. 25,000 Water construction

Water treatment

 $40,000 On site Dredge loader, dump

trucks, bulldozer

1989 Oct. 25,000 Water construction

Water treatment

 $45,000 On site Dredge loader, dump

trucks, bulldozer

Totals 140,000 $252,000

Table 12-10. Chemical Characteristics of Residuals Generated in 1992, Mesa WTP, Colorado Springs, CO (Mesa WTP, 1994)

CITY OF COLORADO SPRINGS WATER QUALITY LABORATORY

SAMPLE REPORT

Sample Site: MPSD Mesa Plant Sediment (MPSD) Date: 03/18/93 09:56 AM

Sample Number: 92120330 Water Type: M

Comments: Sample Date: 12/30/92

Analysis Description/Code Result Result Code

Minimum 

Reported 

Level Test Units Method Code

pH PH 7.03 SU E

Percent solid PS 39.8 mg/L GRAV

Magnesium (dissolved) MGD 3.6 0.10 mg/L FAASA

Silver (dissolved) AGD 0 ND 0.5 μg/L GFAAS

Aluminum (dissolved) ALD 7,300 400 μg/L FAASN

Arsenic (dissolved) ASD 0 ND 4 μg/L GFAAS

Barium (dissolved) BAD 370 100 μg/L FAASN

Cadmium (dissolved) CDD 0.25 0.10 μg/L GFAAS

Chromium (dissolved) CRD 0 ND 50 μg/L FAASN

Copper (dissolved) CUD 0 ND 12 μg/L FAASA

Mercury (dissolved) HGD 0 ND 0.2 μg/L CVAAS

Lead (dissolved) PBD 0 ND 1.0 μg/L GFAAS

Selenium (dissolved) SED 0 ND 1.0 μg/L VGAAS

Key

ND = Not detected.

227
Figure 12-7. Letter from Colorado Springs Department of Utilities, CO (Colorado Springs Utilities, 1994).

228
Figure 12-7. Letter from Colorado Springs Department of Utilities, CO (Colorado Springs Utilities, 1994) (Continued).

229
The water is then filtered through eight double-bay gravity dual media filters. The present permitted filter rate is

6.0 gpm/ft2

. The filters can be operated in the constant

rate or declining rate modes. The filter backwash system

consists of surface sweeps and underdrain backwash.

The filtered water is then treated with lime, fluoride, and

additional chlorine before storage in the plant clearwells.

The finished water is then pumped to customers.

12.3.2 Residuals Handling Facilities

The physical and chemical characteristics of the residuals generated from the two plants differ slightly because

of the different raw water sources and the amount of lime

used by the plants in the dewatering process. The residuals handling facilities for both treatment plants are

the same.

The residuals generated by the two plants each have a

solids content of approximately 35 percent following

pressure filtration. The lime content on a dry weight

basis is approximately 10 to 15 percent. The calcium

carbonate equivalency (CCE) for the material ranges

from 20 to 40 percent. The material is highly plastic until

it is dried further. It has the appearance of moist dirt and

has a slight but noticeable odor.

Extensive testing of the sludge has been conducted.

The Toxicity Characteristic Leaching Procedure (TCLP)

has been used to analyze the residuals from both

plants on several occasions. TCLP analyses for metals,

Figure 12-8. Schematic process diagram of Quarles WTP, Marietta, GA (Cobb County, 1994).

Figure 12-9. Schematic process diagram of Wyckoff WTP, Marietta, GA (Cobb County, 1994).

230
pesticides, herbicides, volatile organics, and semivolatile organics have been conducted. Table 12-12 presents the most recent TCLP data for the Quarles and

Wyckoff plant residuals. The toxics levels have been

extremely low for all samples.

The levels of heavy metals in the sludge have also been

determined and are shown in Table 12-13.

Additional analyses of nutrients and other metals, pesticides and PCBs, triazine herbicides, and reactivity for

cyanide and sulfide were conducted. The data from

these analyses show that this material has a very low

nutrient content. Neither pesticides, PCBs, nor herbicides were detected. The reactivity for cyanide and sulfide was below the EPA criteria limits stated in 40 CFR

Part 261.23. The data for these analyses are reported

in Tables 12-14, 12-15, 12-16 and 12-17.

The Quarles Plant has two gravity thickeners, both 40

feet in diameter with a sidewater depth of 14 feet. The

sludge conditioning takes place in two tanks with

10,000-gallon capacities each. Pebble lime is slaked

and fed to the conditioning tank. Approximately 10 to 15

percent of the dry weight of the dewatered sludge is

lime. The plant controls conditioning by maintaining a

pH of 11.7 to 12.0 in the conditioned sludge. The plant

has two filter presses containing 50 chambers each. The

total volume per press is 60 cubic feet. The plates are

spray washed approximately every load and acid

washed every month.

The dewatered sludge is discharged into dump trucks

and hauled to a sludge storage site at the Wyckoff Plant.

The sludge storage site is an approximately 15-acre

fenced site with gravel roads and a washdown station.

The Wyckoff Plant has three gravity thickeners. Two

units are 40 feet in diameter with a 14-foot sidewater

depth. The third unit is 65 feet in diameter with a 14-foot

sidewater depth. The sludge conditioning tank capacity

is 7,000 gallons. Hydrated lime is mixed with water and

fed to the conditioning tank. Approximately 10 to 15

Table 12-12. TCLP Data, Cobb County-Marietta Water

Authority, Marietta, GA (Cobb County, 1994)

Regulatory

Level Laboratory Data

Compound

EPA

(mg/L)

Quarles

(mg/L)

Wyckoff

(mg/L)

Volatile Organics

Chloroform 6.0 0.013 <0.005

Carbon tetrachloride 0.5 <0.005 <0.005

Tetrachloroethene 0.7 <0.005 <0.005

Clorobenzene 100.0 <0.005 <0.005

1,2-Dichloroethane 0.5 <0.005 <0.005

Benzene 0.5 <0.005 <0.005

Vinyl chloride 0.2 <0.005 <0.005

1,1-Dichloroethene 0.7 <0.005 <0.005

Trichloroethene 0.5 0.016 <0.005

1,4-Dichlorobenzene 7.5 <0.005 <0.005

2-Butanone (MEK) 200.0 <0.10 <0.10

Metals

Arsenic 5.0 <0.005 <0.005

Barium 100.0 1.03 1.22

Cadmium 1.0 <0.005 <0.005

Chromium 5.0 0.12 <0.05

Lead 5.0 <0.05 <0.05

Mercury 0.2 <0.0002 <0.0002

Selenium 1.0 <0.005 <0.005

Silver 5.0 <0.01 <0.01

Pesticides and

Herbicides

Chlordane 0.03 <0.01 <0.01

Endrin 0.02 <0.001 <0.001

Heptachlor 0.008 <0.001 <0.001

Heptachlor epoxide 0.008 <0.001 <0.001

Lindane 0.44 <0.001 <0.001

Methoxychlor 10.00 <0.002 <0.002

Toxaphene 0.50 <0.01 <0.01

2,4-D 10.00 <0.005 <0.005

2,4,5-TP (Silvex) 1.00 <0.005 <0.005

Semivolatile Organics

Cresol (total) 200.0 <0.01 <0.01

1,4-Dichlorobenzene 7.5 <0.01 <0.01

2,4-Ninitrotoluene 0.13 <0.01 <0.01

Hexachlorobenzene 0.13 <0.01 <0.01

Hexachloro-1,3-butadiene 0.5 <0.01 <0.01

Hexachloroethane 3.0 <0.01 <0.01

Nitrobenzene 2.0 <0.01 <0.01

Pentachlorophenol 100.0 <0.05 <0.05

Pyridine 5.0 <0.01 <0.01

2,4,5-Trichlorophenol 400.0 <0.01 <0.01

2,4,6-Trichlorophenol 2.0 <0.01 <0.01

Table 12-13. Total Metals Data, Cobb County-Marietta Water

Authority, Marietta, GA (Cobb County, 1994)

Laboratory Data

Compound

Quarles

(mg/kg)

Wyckoff

(mg/kg)

Arsenic 3.37 3.26

Cadmium 3.37 3.26

Chromium 25 20

Copper 29 35

Lead <5 <5 

Mercury <0.01 <0.01

Molybdenum 5 9

Nickel 15 15

Selenium 0.75 0.44

Zinc 53 58

231
percent of the dry weight of the dewatered sludge is

lime. The plant controls conditioning by maintaining a

pH of 11.5 to 12.0 in the conditioned sludge. The plant

has two filter presses containing 61 chambers, with a

total volume per press of 74 cubic feet. The plates are

spray washed when necessary and acid washed once

per month.

The dewatered sludge is also discharged into dump

trucks and taken to the sludge storage site adjacent to

the Wyckoff Plant.

12.3.3 Ultimate Disposal—Land Application

The two plants together produce approximately 8,000

cubic yards of dewatered sludge per year. The solids

content varies from 25 to 40 percent solids but is usually

around 35 percent solids. Because of the size of the

storage site it is necessary to haul sludge to the application sites two or three times per year.

The ultimate disposal is handled by a contractor for the

Water Authority. The contractor uses trailers with a

dumping capability and that haul 30 cubic yards per

load. Usually, approximately 3,000 to 5,000 cubic yards

of sludge are hauled during an application period, involving up to 170 truck loads.

The Authority has also contracted with a firm to obtain

the land necessary for the application program and to

coordinate the monitoring program. The firm has acquired under contract approximately 1,560 acres to receive the WTP residuals, which are called lime byproduct. As of this writing, approximately nine sites totaling

1,107 acres have received lime byproduct. The firm that

handles the land application program meets with farmers who may be interested in receiving lime byproduct

and discusses the program with them. The farmers’

fields are sampled to determine their lime requirements.

Fields that are included in the program must be of a

certain size and slope to be acceptable. The farmers

must also practice good soil management techniques in

the operation of their farms.

Table 12-14. Other Analyses, Cobb County-Marietta Water

Authority, Marietta, GA (Cobb County, 1994)

Laboratory Data

Compound

Quarles

(mg/L)

Wyckoff

(mg/L)

TKN nitrogen 2,000 5,600

Nitrate nitrogen 300 750

Ammonia nitrogen 100 300

Phosphorus 900 2,000

Potassium 700 500

Sulfur 3,400 2,300

Calcium 161,000 228,000

Magnesium 4,800 6,000

Sodium 100 200

Iron 14,600 18,600

Aluminum 46,000 67,300

Manganese 2,100 2,260

Calcium carbonate equivalency 250,600 338,500

Chloride 700 12,500

Boron 5 15

TOC 66,000 146,000

Table 12-15. Pesticides and PCBs (Solids), Cobb

County-Marietta Water Authority, Marietta, GA

(Cobb County, 1994)

Laboratory Data

Compound

Quarles

(mg/kg)

Wyckoff

(mg/kg)

Aldrin <5.0 <5.0

alpha-BHC <5.0 <5.0

beta-BHC <5.0 <5.0

delta-BHC <5.0 <5.0

gamma-BHC (Lindane) <5.0 <5.0

Chlordane (Technical) <5.0 <5.0

4,4′-DDD <5.0 <5.0

4,4′-DDE <5.0 <5.0

4,4′-DDT <5.0 <5.0

Dieldrin <5.0 <5.0

Endosulfan I <5.0 <5.0

Endosulfan II <5.0 <5.0

Endosulfan sulfate <5.0 <5.0

Endrin <5.0 <5.0

Endrin aldehyde <5.0 <5.0

Heptachlor <5.0 <5.0

Heptachlor epoxide <5.0 <5.0

Mothoxychlor <5.0 <5.0

Mirex <5.0 <5.0

Parathion <5.0 <5.0

Toxaphene <50.0 <50.0

PCB 1016 <20.0 <20.0

PCB 1221 <20.0 <20.0

PCB 1232 <20.0 <20.0

PCB 1242 <20.0 <20.0

PCB 1248 <20.0 <20.0

PCB 1254 <20.0 <20.0

PCB 1260 <20.0 <20.0

232
Actual land application involves transporting the dewatered sludge to a farmer’s field and discharging it in a

location mutually agreeable to the farmer and the contractor. The sludge is then placed into a manure

spreader using a front end loader. The manure spreader

is a side-discharging unit that evenly spreads the material across the field.

The scheduling of the land application is determined by the

farmer involved. The farmer’s planting and harvesting

periods, condition of his fields (particularly the soil dryness), and the availability of the lime byproduct are

considerations. Because of the demand for the material,

some farmers have had to wait a year before they

receive any material.

Lime byproduct is applied to both pasture and cropland.

Approximately 88 percent of the land under contract is

pasture, with the remainder in crops. No restrictions

exist on the types of crops or pasture where the material

is applied, except for the agronomic pH requirements of

the plants involved. Because the land used in the program is not a dedicated site, and because the metal and

nutrient content of the material is very low, there are no

concerns about heavy metal or nitrogen loadings.

The application rate for the material depends on the lime

requirements of the field and the lime content of the

plant residuals. The wet tons per acre value ranges from

approximately 34 to 100 metric tons per hectare (mt/ha)

(15 to 45 tons/acre) and dry tons per acre ranges from

11 to 34 mt/ha (5 to 15 tons/acre). The CCE application

rate ranges from approximately 2 to 5.6 mt/ha (1 to 2.5

tons/acre). As of this writing, it has not been determined

how frequently lime byproduct can be applied on the

same field, but it will probably be no sooner than every

three years. One farmer in the program had material

applied in the fall of 1990 and still had good pH levels

in his fields in the spring of 1993. The material is normally surface applied with no soil incorporation. On

some farms a drag has been run over the field after

application to further break up the material and speed

up its incorporation into the soil. Normally, the material

is washed by rain into the soil within a few weeks.

The monitoring program that the Authority uses for its

land application program includes sampling and analysis of the residuals and the affected soil and vegetation.

Ground-water monitoring is not conducted because the

program does not use a dedicated site. Stream monitoring is conducted on the stream in the drainage area

where the sludge storage yard is located.

The sludge is sampled quarterly for percent solids, percent CCE, pH, and TCLP for certain metals. Semiannual

sampling is conducted for TCLP for a wider list of metals,

total metals, and nutrients. Annual sampling is conducted for chlorotriazine herbicides. TCLP for organics

is presently sampled every other year pending the results of the analyses. PCBs, total organic carbon (TOC),

total Kjeldahl nitrogen (TKN), phosphorus, ammonia nitrogen, and nitrate nitrogen have all been sampled once.

Soil monitoring is conducted on ten fields per year including one field on each farm in the program for that

year. Each field is sampled before and after application

of residuals. Among the analyses conducted are total

metals and standard soil fertility tests. The soil analysis

parameters for the metals tests are: aluminum, arsenic,

boron, cadmium, calcium, chromium, chloride, copper,

iron, lead, magnesium, manganese, mercury, molybdenum, nickel, potassium, selenium, sodium, sulfur, and

zinc. The standard soil fertility test includes tests for:

percent organic matter, nitrate, phosphorus, potassium,

calcium, magnesium, cation exchange capacity (CEC),

pH, and percent base saturation. Soil monitoring was

conducted in a 1991 pilot study of three soil plots. One

plot had commercial lime applied, one had lime byproduct applied, and one had no lime applied. Except for

the pH and calcium content of each plot, no significant

differences were noted between them. Table 12-18 presents the soil data from that study.

Table 12-16. Triazine Herbicides, Cobb County-Marietta Water

Authority, Marietta, GA (Cobb County, 1994)

Laboratory Data

Compound

Quarles

(mg/kg)

Wyckoff

(mg/kg)

Ametryn <1.0 <1.0

Atraton <1.0 <1.0

Atrazine <1.0 <1.0

Cyanazine <1.0 <1.0

Dipropetryn <1.0 <1.0

Prometon <1.0 <1.0

Prometryn <1.0 <1.0

Propazine <1.0 <1.0

Simazine <1.0 <1.0

Simetryne <1.0 <1.0

Terbuthylazine <1.0 <1.0

Tertbutryn <1.0 <1.0

Trietazine <1.0 <1.0

Table 12-17. Reactivity, Solids, Cobb County-Marietta Water

Authority, Marietta, GA (Cobb County, 1994)

Laboratory Data

Parameter

Quarles

(mg/kg)

Wyckoff

(mg/kg)

Reactivity, cyanide <0.25 <0.25

Reactivity, sulfide <5.0 <5.0 

233
Plant tissue analyses will be conducted on several

fields during the first year of the monitoring program. At

least one field on each farm in the program during that

year will be sampled. Row crops are sampled before

harvest, and pastures, or forages, are sampled prior to

grass dormancy. Standard plant analyses will be conducted on the samples. The parameters for these analyses are as follows: aluminum, boron, cadmium, calcium,

chloride, copper, iron, lead, magnesium, manganese,

molybdenum, nickel, nitrogen, phosphorus, potassium,

sodium, sulfur, and zinc.

Plant tissue analyses were conducted on corn grown on

the same three soil plots from the aforementioned 1991

pilot study. No significant differences were found in the

tissue analyses conducted or in the crop yields on the

plots. Table 12-19 presents the plant tissue data from

that study. The average yields for the plots ranged from

122.9 bushels/acre to 130.9 bushels/acre with the lime

byproduct-treated plot having the highest yield.

Stream monitoring is conducted quarterly on sampling

sites above and below the lime byproduct storage area

at the Wyckoff Plant. The analyses evaluate pH, calcium, aluminum, and turbidity levels. The data at present

show little variation between the upstream and downstream sites.

12.3.4 Disposal Costs

The costs for dewatering residuals include personnel,

equipment, chemical, and maintenance costs. The total

1992 costs for Quarles and Wyckoff plants were approximately $360,000. These costs do not include employee benefits or the costs for capital improvement

projects.

The costs for the land application program are on a

cost-per-ton basis, excluding laboratory costs. The elements of the program include land acquisition, transportation to the application site, spreading, monitoring

support, and annual reporting. During 1992, this program cost both plants a total of approximately $170,000.

The laboratory monitoring costs for the first year of the

CCMWA’s program will be about $10,000. In subsequent years the Authority will do many of the analyses

themselves and eliminate others that have not been

detected in the testing. Based on this reduced program,

external laboratory costs should be reduced to about

$2,000 per year.

Table 12-19. Pilot Study, Plant Tissue Data, Cobb

County-Marietta Water Authority, Marietta, GA

(Cobb County, 1994)

Component (ppm)

Plant

Tissue

Without

Sludge

Plant

Tissue

With

Agricultural 

Limea

Plant

Tissue

With

Sludge

Nitrogen 22,900 18,400 23,200

Phosphorous 3,100 3,200 3,800

Potassium 20,000 13,250 20,300

Aluminum 80 30 90

Cadmium <0.5 <0.5 <0.5

Chromium <5 <5 <5

Copper 9 8 9

Nickel <5 <5 <5

Lead <5 <5 <5

Calcium 2,900 2,900 3,500

Magnesium 1,800 1,800 1,800

Boron 8 7 9

Iron 106 96 119

Manganese 68 80 60

Molybdenum 0.4 0.9 0.6

Sodium 100 100 100

Zinc 22 25 22

a Only four samples were taken on these plots compared with six

samples on the other plots. This sampling difference accounts for

lower levels for nitrogen, potassium, aluminum, and calcium.

Table 12-18. Pilot Study, Soil Data, Cobb County-Marietta

Water Authority, Marietta, GA (Cobb County,

1994)

Component (ppm)

Soil

Without

Sludge

Soil With

Agricultural

Limea Soil With

Sludge

pH 5.7 6.6 6.7

Nitrogen 22,900 18,400 23,150

Phosphorous 3,100 3,200 3,750

Potassium 24,900 13,300 20,280

Aluminum 80 30 90

Cadmium <0.5 <0.5 <0.5

Chromium <5 <5 <5

Copper 9 8 9

Nickel <5 <5 <5

Lead <5 <5 <5

Calcium 2,900 2,875 3,500

Magnesium 1,800 1,800 1,800

Boron 8 7 9

Iron 106 96 119

Manganese 68 80 60

Molybdenum 0.4 0.9 0.6

Sodium 100 100 100

Zinc 22 25 22

a Only four samples were taken on these plots compared with six

samples on the other plots. This sampling difference accounts for

lower levels for nitrogen, potassium, aluminum, and calcium.

234
The economic benefits gained from the land application

program are based on the avoided costs of tipping fees

that must be paid to the local landfill. These costs are

expected to increase in the years to come with the

increase in regulations and the difficulty in obtaining

landfill sites. Based on landfill tipping fees of $32.50/ton,

the annual cost would be about $220,000 per year.

Because the dewatering costs are incurred whether the

material is land applied or placed in a landfill, the savings per year are approximately $50,000. During the first

year of the present monitoring program the savings will

be less due to the higher monitoring costs.

12.4 Case Study 4: Treatment of

Residuals at the Lake Gaillard Water

Treatment Plant, North Branford,

Connecticut

12.4.1 Facility Information

12.4.1.1 Major Plant Modifications

• Added four new rapid sand dual media filters to the

plant, thereby expanding the number of filters in the

plant to 16, and increasing plant capacity from 60

mgd to 80 mgd.

• Expanded flocculation basins by 20 percent by adding a new flocculation basin system.

• New control valves and magmeters were added to

outflow of equalization basins to optimize the clarifier

loading rate.

• Doubled the number of clarifiers.

• Installed new sludge pumps-rotary lobe (variable frequency drive).

• Installed new recycle pumps and controls (variable

frequency drive).

• Installed new sludge lagoon/freeze-thaw/sand drying

beds.

12.4.1.2 Sludge Lagoon/Freeze-Thaw/Sand

Drying Beds

Lagoon Construction

• Site constraints influenced lagoon construction, i.e.,

wetlands required lagoons to be constructed with

gabions (1 on 1 slopes).

• Site includes approximately 20,000 square feet of

lagoon bottom in each of the four lagoons.

• Lagoons were initially constructed with natural soil

bottoms and gravity decant chambers.

• Following unsatisfactory performance of the first lagoon once placed into service, all lagoons were

retrofitted with a sand/filter fabric/gravel PVC piping

underdrain system.

12.4.1.3 Dewatering Method as Designed

Backwash Water From Equalization Basin to Gravity

Clarifiers

• Clarifiers Parkson (lamella) and EIMCO (stainless

“Flex-Klear”) parallel units.

• Clean water recycled to head of the plant.

• Sludge thickened to approximately 2 percent solids.

Sludge Pumped by Rotary Lobe Pumps to Natural

Silt Bottom/Freeze-Thaw Lagoons

Evaporation

Decant Through Gravity Decant Chambers Back to

Equalization Basins

12.4.1.4 Retrofitted Dewatering Method

• Gravity clarifiers (same).

• Sludge pumped by hose pumps to sand drying

bed/freeze-thaw lagoons with piped underdrain systems. Rotary lobe pumps could not achieve 2 percent

solids based on our piping arrangement, and anthracite particles caused the pumps to fail.

• Evaporation.

• Decant through gravity chambers (limited).

• Gravity drainage through underdrain system.

• Some initial short circuiting was noted at the interior

lagoon perimeter. Modifications were made to the

filter fabric/sand/gabion interface that solved this

problem. This type of short circuiting (at any vertical

interior lagoon boundary condition) is something that

needs to be considered very carefully during lagoon

design and construction.

• Installed vertical filter drains to enhance dewatering

from sludge column to sand underdrain system.

• Conducted summer furrowing of thin sludge layers to

increase evaporation by breaking up the top skim

layer, which greatly increased the effective surface

area of the sludge.

• Conducted mechanical (in-lagoon) winter mixing of

the sludge to enhance the freeze-thaw process. This

method allows freezing of three to inch-inch layers

each night.

• Following the freeze/thaw process, the sludge at approximately 40 percent solids is removed from the

lagoon by front end loader and trucked to a monofill

235
owned and operated by the Lake Gaillard Water

Treatment Plant.

12.4.2 Lessons Learned

• Sludge will not dewater by evaporation and decanting

alone.

– Some means of drainage at the bottoms of lagoons

is required. At this plant a sand/gravel underdrain

system with PVC piping was installed.

– A skim layer forms on the surface of sludge causing a rapid decrease in the rate of evaporation.

– The skim layer should be broken up during the

summer months to enhance evaporation and simultaneously to increase the surface area. Surface

cracking down to the sand layer (while beneficial

to enhance dewatering and channel away rain) is

not in itself sufficient to dewater the sludge. In our

case, furrowing with a small tractor over thin (8 to

10 inches) layers in summer produced good results, i.e., 25 to 30 percent solids.

– Decanting a lagoon while it is being filled with new

sludge may result in short circuiting. This happened at this plant and is unacceptable because

decant water is directed back to the equalization

basins/recycle system, causing undesireable turbidity to the head waters of the WTP.

• Sludge blinds the sand layer at the surface boundary

of the sand but not completely. Blinding does not

occur throughout the layer of sand depth.

– The rate of underdrain flow will decrease over time

given a relatively constant head of sludge in the

lagoon. The headloss is due to a blinding layer

immediately at the sand layer surface boundary.

– What appears to be an anaerobic (dark) layer of

sludge becomes evident at the sand layer surface

boundary. At this plant, the dark layer was approximately one to several inches thick after several

(summer) months of initial application.

– In the case of this plant, what appeared to be an

(orange) iron precipitate formed at the filter fabric

layer between the bottom of the sand layer and

the top of the gravel underdrain pipe layer. Over

time this may blind the filter fabric.

– Vertical filter drains, designed by the plant operator, were found to enhance the dewatering process

by automatically decanting surface water (and

some sludge column), and directing it to the sand

underdrain system.

• Sand layers should be approximately 12 to 18 inches

thick, to protect the filter fabric below the sand layer.

– Heavy equipment entering the lagoon digs into the

sand and tears the fabric if it is too close to the

sand surface. At this plant, the initial sand depth

was 6 inches and we experienced substantial tearing of the filter fabric. After the initial cleaning of

the lagoons the sand layer was increased to 12

inches. Approximately two inches of sand was removed during the lagoon cleaning process.

– Track equipment is superior to rubber tire equipment for working on sand drying beds.

– Careful attention to the sand/filter fabric/gabion

connection at the lagoon interior perimeter needs

to be made during construction to avoid short circuiting of sludge into the underdrain system.

• Lagoons divided into several “compartments” or “sub

lagoons” are beneficial because sludge drying and

cracking do not proceed effectively until loading of

the lagoon is complete. This has not been implemented at Lake Gaillard WTP at this time.

– Loading should be completed in thin layers (three

to four inches) to enhance summer drying or winter

freeze-thawing. Lagoon compartments will accommodate this.

– Sufficient area is the key to the successful operation of sand drying/freeze-thaw beds with underdrains. This plant is seeking ways to expand its

workable lagoon area.

• Only the top three or four inches of sludge will freeze.

This frozen layer then insulates the sludge below and

stops the freezing process.

– Winter mixing by use of a track-mounted backhoe

or other similar equipment can expose fresh

sludge for freezing every night. A layer of approximately three inches can be frozen every night at

temperatures of 20°F or less.

– A thin layer of snow will insulate the sludge lagoon

and prevent the freezing process.

• The sludge lagoons produce a “musty,” “low tide”

odor during the summer months.

• Elevated levels of chloroform were noted in the underdrain water during the spring melt period (1992).

This may be due to prechlorination at the plant.

12.5 Case Study 5: Disposal of Water

Treatment Plant Residuals From the

San Benito Water Plant, Brownsville,

Texas

The San Benito Water Plant is located in the Rio Grande

Valley near Brownsville, Texas. The water plant is a

mixed media, rapid filtration, surface water plant. The

source of water for the San Benito plant is the Rio

Grande River. Water from this river is pumped into a

system of canal-type lakes called Resacas. The design

capacity of the plant is 6 mgd. The maximum daily

demand is approximately 2.7 mgd in the winter, and 4

236
mgd in the summer. WTP residuals are removed from

the San Benito plant as follows:

1. Sludge is removed from two upflow clarifier units to

storage lagoons through a drain system at the bottom of the clarifiers. Backwash water (205,000 gallons per day) containing residuals is also directed to

the storage lagoon. The lagoon is equipped with a

pump to return excess water to the headworks, to be

treated again.

2. Settled sludge is retained in the storage lagoons for

approximately 3 months.

3. At the end of the 3-month period, the plant uses a

tractor loader and dump truck to remove the thickened sludge to a local landfill.

12.5.1 Residuals Handling Facilities

The storage lagoons are lined with concrete below

ground level. Sludge is removed from the treatment

plant to the lagoons by gravity flow, thus saving pumping

costs.

The plant does not use any sludge thickening processes. The supernatant is removed after the sludge

settles, and the sludge is exposed to the sun. Approximately 2 weeks before removing the sludge from the

lagoons, a divider is used to separate the basin into two

halves. The portion not receiving flow dries in the hot

Texas sun. The sludge is then dry enough for a tractor

loader to remove it to a landfill.

12.5.2 Final Disposal

The sludge is disposed of in a local landfill that is classified to receive this type of waste. At the time of this

writing, the operation of the landfill had recently been

privatized. Instead of being run by the City of San Benito, it is to be operated by a private company. The San

Benito WTP intends to continue disposing of its residuals at this landfill.

12.5.3 Disposal Costs

Labor costs form the greatest portion of the San Benito

plant’s total disposal costs. Normally, sludge removal to

the landfill requires a week of two people working fulltime. In 1992, labor costs totaled an estimated $5,000.

Equipment costs were estimated at $1,000, including

fuel and maintenance costs. The plant’s estimated total

annual disposal costs for 1992 were $6,000.

12.6 Case Study 6: Management of Water

Treatment Plant Residuals in the

Chicago Area, Chicago, Illinois

The major source of drinking water in the metropolitan

Chicago area is Lake Michigan. Seven water treatment

plants draw water from Lake Michigan to supply drinking

water to Chicago and various neighboring communities.

The largest water system is the City of Chicago system,

which serves a population of more than 4.6 million people in Chicago and 118 suburban municipalities in both

Cook and DuPage counties. The Chicago water system

includes two of the largest water treatment plants in the

world (Jardine Water Purification Plant (WPP) and

South Water Filtration Plant (WFP)), three water intake

cribs located 2 to 2.5 miles offshore in Lake Michigan,

11 pumping stations, and thousands of miles of water

mains. The finished water from the Chicago water system is supplied for drinking and industrial use.

Six other municipalities along the lakeshore (Evanston,

Glencoe, Kennilworth, Northbrook, Wilmette, and Winnetka) obtain water directly from Lake Michigan and

maintain their own treatment plants, which serve a combined population of approximately 0.44 million people.

Twenty-five years ago, the residuals generated from all

of the abovementioned WTPs were disposed of by direct discharge into Lake Michigan. In keeping with a

mandate to prevent pollution of Lake Michigan, the Metropolitan Water Reclamation District of Greater Chicago

(the District) adopted an ordinance that contained a

provision prohibiting the discharge of waste of any kind

to Lake Michigan. In the early 1970s, the District took

action requiring WTPs to eliminate the practice of discharging their residuals into Lake Michigan. By 1974, all

of these discharges were diverted into the sewerage

system of the District.

The District covers over 872 square miles within Cook

County, and serves 125 municipalities including the City

of Chicago. The population served by the District is 5.2

million people. In addition, a nondomestic equivalent of

4.5 million people is also served by the District, resulting

in a total sewage treatment load equivalent to almost 10

million. The wastewater collection system operated by

the District consists of more than 500 miles of intercepting sewers and 23 pumps stations. The District owns

and operates seven water reclamation plants (WRPs)

with a combined maximum capacity of about 2,000 million gallons. They include the world’s largest activated

sludge plant, the Stickney WRP, which has a maximum

capacity of 1,440 mgd. All of the other District WRPs are

also activated sludge treatment plants. About 500 dry

tons of sludge solids are generated per day from these

plants.

This case study focuses on the Jardine WPP, which

discharges its residuals into the District’s Stickney WRP.

12.6.1 Description of the Jardine Water

Purification Plant

The Jardine WPP is located in Chicago on Lake Michigan, just north of the Chicago River. It has a design

237
capacity of 960 mgd based on a sand filter design rate

of 2 gpm/ft2

. A schematic flow diagram is shown in

Figure 12-10. The plant has 192 sand filters, each with

a nominal filter rate of 5 mgd. Water is drawn from Lake

Michigan either through a 20-foot diameter tunnel from

a crib located 2.5 miles out in the lake, or directly

through a shore intake at the plant.

The unit processes at the Chicago WTPs include traveling screens to remove fish, aquatic weeds, and trash.

The water from Lake Michigan is then pumped upward

21 feet to permit gravity flow through the WFP. Chlorine,

fluoride, and polyelectrolyte are added after the screen

in the intake basin.

Alum is added ahead of flocculation basins where slowspeed mixers enhance the flocculation process. The

flocculation water then passes into settling basins with

a detention time of 4 hours. The settled water is then

conveyed to sand filters. Hydrated lime is added ahead

of the filters to minimize corrosion in the water distribution system. After filtration, the water flows to filtered

water clearwells where additional chlorine is added,

along with caustic soda. From the clearwells the water

is taken to a 10-acre filtered water reservoir prior to

being conveyed to the distribution system.

Sludge from the settling basins of the WFP is removed

and collected in a holding tank. The filter backwash is

returned to the settling basins. Organic material removed from the intake screens is macerated and

pumped into the sludge holding tank. The collected

sludge in the holding tank is pumped at a controlled rate

into an interceptor sewer of the District. A schematic of

the sludge management system at the Jardine WPP is

presented in Figure 12-11.

12.6.2 Residuals Handling and Disposal

Table 12-20 shows the volume and strength of the water

treatment residuals discharged by Chicago’s Jardine

and South WFP into the District’s WRPs from 1984 to

1992. The Jardine WPP’s residuals volume ranged from

192 to 378 million gallons per year, with suspended

solids ranging from 7,605 to 11,148 tons per year (20.1

to 30.6 dt/day). These quantities are roughly 5 to 6

percent of the residuals produced by the District’s

WRPs. The biochemical oxygen demand (BOD) is relatively low, less than 200,000 pounds per year, because

the majority of the sludge solids is inorganic. Similar

data for the other six WFPs, which treat Lake Michigan

water and discharge their sludges into the District system, are summarized in Table 12-21.

The District has, under its Sewage and Industrial Waste

Ordinance, imposed local limits for several heavy metals

(see Table 12-22). A summary of the metals concentrations observed in the WTP residuals is shown in Table

12-23. A comparison of the residual metals concentrations with the limits imposed by the Ordinance indicates

Figure 12-10. Schematic flow diagram of Jardine Water Purification Plant, Chicago, IL (MWRDGC, 1994).

238
that the residuals metal concentrations are within the

Ordinance limits, except for lead at the Evanston, Winnetka, and Wilmette plants.

12.6.3 Disposal Costs

The costs for discharging wastes into the District’s treatment system are based on volume, BOD, and suspended solids. In 1992, the user charge rates for

tax-exempt entities such as the WFPs were:

Volume $131.60 per million gallons

BOD $148.34 per thousand pounds

Suspended solids $132.74 per thousand pounds

A summary of the annual charges paid by the City of

Chicago’s two WTPs for the period of 1984 through

1992 is presented in Table 12-24. In 1992, the Jardine

WPP and South WFP paid $2,723,317 and $1,726,593

in user charges, respectively. In 1990, the Jardine WPP

paid $1,837,940, or an average of $5,035 per day, to

annually dispose of 378 million gallons of residuals containing 16.4 million pounds of suspended solids (approximately 22.5 tons per day). During 1990, the WTP

supplied 214 billion gallons of filtered water to its consumers. From 1984 through 1992, the average cost to

the City of Chicago for disposal of the WTP residuals

was $0.00858 per thousand gallons of finished water, or

$224 per ton of dry solids.

For the years 1984 through 1992 the residuals disposal

costs paid by the six water treatment plants on Lake

Michigan not operated by the City of Chicago are summarized in Table 12-25. The total costs are based upon

volume and BOD as well as solids; thus, the variation in

Figure 12-11. Schematic diagram of residuals management system, Jardine Water Purification Plant, Chicago, IL (MWRDGC, 1994).

Table 12-20. Water Treatment Sludges Discharged by the City of Chicago WTPs to the District, Chicago, IL, 1984–1992

(MWRDGC, 1994)

Jardine WPP Plant South WFP

Year

Volume

(MG) BOD (lb) TSS (lb x 103

)

Volume

(MG) BOD (lb) TSS (lb x 103)

1984 205.66 20,441 22,296 355.9 117,955 10,956

1985 215.53 226,447 21,118 336.5 179,637 12,036

1986 191.59 182,156 18,405 396.3 185,074 8,920

1987 200.82 166,822 17,869 471.3 146,008 13,069

1988 176.75 271,268 15,209 339.6 185,567 11,222

1989 244.65 272,126 17,588 315.1 131,116 11,570

1990 378.26 276,049 16,417 323.9 165,435 12,913

1991 362.87 238,360 18,439 326.5 119,815 11,486

1992 239.1 305,342 19,937 382.3 95,174 12,522

239
Table 12-21. Water Treatment Sludges Discharged by Various Suburban WTPs to the District, Chicago, IL, 1984–1992

(MWRDGC, 1994)

Evanston Wilmette Northbrook

Year

Volume

(MG)

BOD

(lb)

TSS

(lb x 103

)

Volume 

(MG)

BOD

(lb)

TSS

 (lb x 103)

Volume

(MG)

BOD

(lb)

TSS

 (lb x 103)

1984 2.056 7,928 528 18.06 4,497 679 3.77 5,320 353

1985 2.032 14,436 456 13.661 4,494 705 3.77 3,716 340

1986 1.286 3,652 364 15.406 11,716 755 2.407 3,212 363

1987 1.601 5,480 851 11.055 3,436 788 1.414 2,521 334

1988 0.898 3,383 466 12.629 5,918 685 0.777 1,962 176

1989 7.335 28,996 2,447 12.740 6,418 597 0.976 1,703 183

1990 6.345 26,547 2,238 11.326 5,671 565 1.709 4,396 481

1991 3.801 17,109 2,350 13.034 6,957 820 1.889 3,386 512

1992 8.521 31,711 2,641 11.244 3,881 644

Winnetka Kennilworth Glencoe

Year

Volume

(MG)

BOD

(lb)

TSS 

(lb x 103

)

Volume 

(MG)

BOD

(lb)

TSS

(lb x 103)

Volume

(MG)

BOD

(lb)

TSS 

(lb x 103)

1984 0.683 1,563 330 0.44 60 33 0.324 354 102

1985 0.631 1,761 220 0.44 171 22 0.342 661 52

1986 1.694 2,057 321 0.44 408 22 0.388 618 58

1987 1.361 1,879 198 0.44 200 24 0.398 915 129

1988 1.096 3,678 271 0.44 270 86 0.359 777 47

1989 1.111 1,712 114 0.44 220 34 0.337 744 63

1990 1.423 4,161 246 0.44 387 45 0.238 835 97

1991 1.075 5,444 308 0.44 288 55 0.292 613 108

1992

Table 12-22. Local Limits on Dischargers Into District

Sewarage Systems, Chicago, ILa

(MWRDGC, 1994)

Pollutant

Maximum

Concentration

(mg/L)

Cadmium 2.0

Chromium (total) 25.0

Chromium (hexavalent) 10.0

Copper 3.0

Cyanide (total) 5.0

Fats, oils, and greases (FOG) 250.0

Ironb 250.0

Lead 0.5

Nickel 10.0

Zinc 15.0

pH range: not lower than 5.0 or greater than 10.0

a Sewage and Waste Control Ordinance as amended September 5,

1991. b Discharges from domestic WTPs that supply potable water to the

general public are exempt from this limitation for iron.

Table 12-23. Concentrations of Metals Found in WTP

Sludges Discharged Into the District, Chicago,

IL (MWRDGC, 1994)

Concentration in mg/L

Total

Water

Treatment Plant Cd Cu Cr Fe Ni Ph Hga Zn

Jardine 0.01 0.02 0.04 7.1 0.0 0.03 0.3 0.1

South (Chicago) 0.02 0.08 0.06 11.1 0.1 0.07 0.9 0.2

Evanston 0.29 1.40 1.44 866 1.3 4.89 1.5 5.2

Wilmette 0.18 0.58 0.83 225 0.7 2.37 0.7 2.6

Winnetka 0.23 1.48 0.90 464 0.9 4.35 2.6 2.8

Northbrook 0.03 0.06 0.05 8.5 0.1 0.09 0.1 0.1

a Hg concentration in μg/L.

240
the costs per dry ton is due to differences in the amount

of water discharged with the solids. The average annual

cost of discharging WTP residuals from the Chicago

facilities ranged from $159/dt in 1984 to $275/dt in 1992

(see Table 12-24). Similarly, the average annual cost for

discharging sludges from the other treatment plants

(see Table 12-25) ranged from $164/dt in 1984 to

$246/dt in 1991.

Table 12-24. User Charge Costs for Disposal of Water

Treatment Residuals to the Metropolitan Water

Reclamation District, Chicago, IL, 1984–1992

(MWRDGC, 1994)

Year

Jardine

WRP Cost/dt South WFP Cost/dt

Average

Cost/dt

1984 $1,744,204 $156 $887,063 $162 $159

1985 $1,617,996 $153 $944,780 $157 $155

1986 $1,317,075 $143 $669,625 $150 $147

1987 $1,511,031 $169 $1,135,339 $174 $172

1988 $1,188,896 $156 $893,417 $159 $158

1989 $1,762,091 $200 $1,167,141 $202 $201

1990 $1,837,940 $224 $1,442,383 $223 $224

1991 $2,335,947 $253 $1,464,025 $255 $254

1992 $2,723,317 $273 $1,726,593 $276 $275

Key

dt = dry ton.

Table 12-25. Water Treatment Sludges Discharged From Various Water Treatment Plants to the District From 1984 Through 1992

(MWRDGC, 1994)

Evanston Wilmette Northbrook Winnetka Kennilworth Glencoe

Year

Volume 

(MG)

BOD

(lb)

TSS

(lb x

103)

Volume 

(MG)

BOD

(lb)

TSS

(lb x

103)

Volume 

(MG)

BOD

(lb)

TSS

(lb x

103)

Volume 

(MG)

BOD

(lb)

TSS

(lb x

103)

Volume 

(MG)

BOD

(lb)

TSS

(lb x

103)

Volume 

(MG)

BOD

(lb)

TSS

(lb x

103)

1984 2.056 7,928 528 18.06 4,497 679 3.77 5,120 353 0.683 1,563 330 0.44 60 33 0.324 354 102

1985 2.032 14,436 456 13.661 4,494 705 3.77 3,716 340 0.631 1,761 220 0.44 171 22 0.342 661 52

1986 1.286 3,652 364 15.406 11,716 755 2.407 3,212 363 1.694 2,057 321 0.44 408 22 0.388 618 58

1987 1.601 5,480 851 11.055 3,436 788 1.414 2,521 334 1.361 1,879 198 0.44 200 24 0.398 915 129

1988 0.898 3,383 466 12.629 5,918 685 0.777 1,962 176 1.096 3,678 271 0.44 270 86 0.359 777 47

1989 7.335 28,996 2,447 12.740 6,418 597 0.976 1,703 183 1.111 1,712 114 0.44 220 34 0.337 744 63

1990 6.345 26,547 2,238 11.326 5,671 565 1.709 4,396 481 1.423 4,161 246 0.44 387 45 0.238 835 97

1991 3.801 17,109 2,350 13.034 6,957 820 1.889 3,386 512 1.075 5,444 308 0.44 288 55 0.292 613 108

1992 8.521 31,711 2,641 11.244 3,881 644

241
Chapter 13

Waste Minimization and Reuse

In the process of treating a raw water supply for potable

water use, contaminants are removed from the raw

water typically in the form of residual solids, brines, and

concentrates. The treatment and disposal of these

waste streams have been discussed extensively in prior

chapters. Historically, these waste streams have been

stored, treated, and then disposed of.

Many years ago, the typical method of disposal was

direct discharge into a nearby receiving stream, or lagoon storage. Often, solids would accumulate in an

onsite lagoon for years. Once capacity was reached, the

waste stream would simply flow directly into an overflow

structure, onto adjacent properties, or into receiving

streams, with little or no settling. This would typically

leave the water utility with an immediate problem, requiring removal of solids and, often, onsite disposal with few

or no control measures. The potential for polluting the

environment, both aquatic and terrestrial, was high.

Over the years, there has been an increased emphasis

on minimizing negative impacts on the environment

through pollution prevention and reuse of materials. This

is most evident in the adoption of the National Pollutant

Discharge Elimination System (NPDES) program. The

implementation of the NPDES program has eliminated

direct discharge to a receiving stream as a viable option

for many water supply systems. Uncontrolled onsite

disposal is generally no longer permissible, requiring the

utility to construct a monofill, discharge to a sanitary

sewer system, or transport the material to a nearby sanitary landfill.

While all of these are viable options, implementation is

becoming more problematic. Construction of an onsite

monofill requires available land and is costly both to

construct and to monitor. Discharge to a sanitary sewer

is often limited by a wastewater treatment plant’s capacity to accept more flow and inert solids, both of which

might affect the wastewater treatment process. 

The regulatory framework that controls the wastewater

industry has also become increasingly stringent, requiring higher levels of wastewater treatment and increasing

costs. In addition, the impact of 40 CFR Part 503 on the

wastewater industry has caused increased concerns

over accepting new material into the waste stream that

may be perceived to limit disposal or reuse options.

Transport and disposal to a nearby landfill is also increasingly difficult. Many landfills have reached capacity

or are nearing capacity, requiring utilities to look beyond

the immediate area for disposal options and significantly

increasing costs. Often, the associated costs of disposal

drive the utility to consider means of reducing quantities

of residuals generated and to give more consideration

to recovery or reuse options.

The net result of these changes is that utilities need to

develop long-term, flexible management plans. They

need to consider innovative techniques for reducing the

quantities of waste, and they need to implement reuse

options. Ultimately, this will result in more cost-effective

solutions for residuals handling and will improve public

acceptance. The dilemma, however, is that some of the

treatment techniques for reducing residual waste stream

volume or recovering chemical aids (i.e., coagulant or

lime recovery) have had limited application over the

years and, at present, may be costly. The markets for

beneficial reuse of residuals at this time are limited.

Some of the available waste minimization, recovery, and

reuse options that apply primarily to residuals generated

from the coagulation/filtration or lime softening processes are discussed below.

13.1 Waste Minimization

13.1.1 Process Modifications

Perhaps the best means of reducing waste is to minimize or optimize the types and quantities of coagulants

and chemical aids used and to optimize associated

mixing conditions. Control at the source is the most

effective first step and should be evaluated in all facilities. Process optimization allows facilities to achieve a

balance between water quality, chemical addition, and

physical parameters. Ultimately, the quality of the raw

water supply will dictate the type of treatment provided.

For example, EPA’s proposed Disinfection/Disinfectant

Byproduct Rule may require utilities to increase coagulant dosages to maximize water quality benefits and

reduce the potential for disinfectant byproduct formation.

242
13.1.2 Dewatering

Dewatering—both mechanical and nonmechanical—is

a potential step in the process of minimizing waste.

Details of this process are provided in Chapter 4.

Several mechanical dewatering techniques are available and employed throughout the country. These typically include filter presses and centrifuges. Mechanical

dewatering significantly reduces the residuals volume

by removing some of the water, thereby increasing the

percent solids content. Several other mechanical dewatering processes have the potential for increased use in

the future. As discussed by Malmrose and Wolfe (1994)

and presented in Slib, Schlamm, Sludge (Cornwell and

Koppers, 1990), these include:

• Mechanical freeze-thaw machine

• Centrifuge

• Diaphragm filter press

• Continuous high-pressure (CHP) filter press

• Tubular filter press

• Cross-flow microfiltration process

• Hi-Compact press

• Compactor

The applicability of these processes depends on raw

water quality, coagulant practices, and desired dewatered cake characteristics. The need for chemical conditioning or pretreatment is also site specific. Most of

these techniques were developed for application in the

wastewater industry. As a result, little information is

available relative to performance in the drinking water

industry. A brief description of these alternatives is provided below.

The mechanical freeze-thaw machine employs freezing

as the method of separating the solid and liquid phases.

This method is employed primarily in Europe; no operating facilities exist in the United States. Energyintensive and generally costly, the process is initiated

by freezing through refrigeration, followed by a thawing

process. This conditioning process, when followed by

other dewatering processes, can achieve a solids content of 30 to 50 percent (Malmrose and Wolfe, 1994).

Scroll and spinning centrifuges have been used in the

United States for many years; however, they typically

have been associated with wastewater operations.

Solid-liquid separation occurs through application of

centrifugal force. Application in the drinking water industry has typically yielded a dewatered cake with a solids

content on the order of 15 percent. Recent improvements in the equipment, as well as the availability of

more operational data, indicate that the dewatered cake

can achieve a solids content in excess of 20 percent,

depending on the solids characteristics, velocity gradient (g value), and feed concentrations. Thickening to 2

to 4 percent prior to centrifugation yields improved results. Centrifuges are usually lower in cost, smaller in

size, and easier to operate than many other types of

mechanical dewatering equipment.

The diaphragm filter press is similar to a basic filter

press, which comprises filter plates that are pressed

together to form a chamber. The plates are lined with a

porous cloth. Residuals are pumped into the chamber;

the solids are retained by the cloth liner, and the liquid

fraction is pushed through as the driving pressure increases with continued pumping. In the diaphragm

press, the liner is replaced with a membrane which,

when pressurized, aids the dewatering process by

squeezing the solids retained in the chamber. This batch

process can achieve a dewatered cake with a solids

content in excess of 30 percent (Cornwell and Koppers,

1990).

The CHP filter press combines elements of both the belt

and pressure filters. An initial gravity thickening zone is

followed by wedging the residuals between two belts at

a low pressure (35 psi). High pressure, on the order of

300 psi, is then applied through cylinders that exert

pressure to a mobile armored belt (Malmrose and Wolfe,

1994).

The tubular filter press and cross-flow microfiltration

process feed residuals, under pressure, to a tubular

matrix. The tubes are composed of porous material, and

the solids are retained on the tubes. In the tubular filter

press, the solids are removed by rollers; the cross-flow

microfiltration unit employs a vortex cleaning system

(Malmrose and Wolfe, 1994). Conditioning prior to application on the presses may not be required. Limited

experience indicates that a solids content in the 20 to

40 percent range can be achieved (Malmrose and

Wolfe, 1994; Cornwell and Koppers, 1990).

The remaining processes, specifically the Hi-Compact

press and the Compactor system, have limited full-scale

applications, even in the wastewater industry. The HiCompact press requires prior thickening and dewatering

to greater than 20 percent solids content.

13.1.3 Drying

As discussed in Chapter 4, a drying step would serve to

further reduce the quantities of solids generated for

reuse or disposal. Two processes for drying are the

Carver-Greenfield and electroacoustical dewatering

(EAD) processes. They are difficult to operate and potentially very costly because of elaborate equipment and

high energy requirements. The Carver-Greenfield process, however, may be able to achieve a dewatered cake

with a greater than 90 percent solids content, thereby

reducing the volume of residuals significantly. Applica243
bility to the drinking water industry is still somewhat

unclear, however. Continued investigations are warranted.

13.2 Chemical Recovery

13.2.1 Coagulant Recovery

Coagulant recovery provides a method for recovering a

resource (i.e., coagulant) and for minimizing waste by

extracting aluminum or iron coagulants from the waste

stream. Extraction is achieved by acidification, which

puts the metals back into solution. Critical design and

operations factors include extraction pH and acid contact time. Extraction pH is typically in the range of 1.8 to

3.0. Acid contact time of 10 to 20 minutes appears to be

reasonable based on full-scale operations data and

laboratory testing (Saunders and Roeder, 1991). Another critical component is the amount of metal coming

from the raw water compared with the coagulant. For

example, aluminum associated with raw water solids is

more difficult to dissolve than that associated with an

alum coagulant.

The quality of the recovered coagulant is an important

factor in determining the feasibility of coagulant recovery. The impact of coagulant reuse on treatment plant

operation and resulting finished water quality must be

given serious consideration. While the acidification process is beneficial for achieving dissolution of the coagulant, concentrations of coagulant impurities and raw

water contaminants may also become dissolved and get

recycled to the head of the plant. While these contaminants may again be removed from the water supply by

the coagulant, settling, and filtration processes, the net

result is a concentration of contaminants. This impact

must be carefully considered as the regulatory climate

becomes increasingly stringent.

Concentration and recycling of contaminants may present a problem in achieving compliance. Of particular

interest is the recycling of organic material, which may

increase disinfection byproduct formation potential. This

may be due in part to the demonstrated need for increased chlorination when organic matter is recycled

(Saunders and Roeder, 1991). Full-scale operations experience at the Vernon S. Wade plant in Athens, Georgia, also indicates an increased alkalinity demand as a

result of using the recovered coagulant.

Coagulant recovery has been investigated in the United

States over the last 30 to 40 years; however, few fullscale facilities are currently active (Saunders and

Roeder, 1991). Saunders and Roeder (1991) have reviewed coagulant recovery extensively and present their

findings in the American Water Works Association Research Foundation research report Coagulant Review:

A Critical Assessment. Pilot- or laboratory-scale operations should be performed when evaluating the feasibility of implementing this option.

13.2.2 Lime Recovery

Lime recovery is accomplished through the recalcination

process whereby the lime residual is dewatered and

then burned to produce quicklime (Cornwell et al.,

1987). As with other recovery processes, strong consideration must be given to the expected quality of the

recovered product and to incorporating processes to

remove impurities. For example, as presented in Handbook of Practice: Water Treatment Plant Waste Management (Cornwell et al., 1987), the first step is

generally a purification process such as centrifugation.

After purification and dewatering of the calcium carbonate, a drying process is generally employed. A variety of

furnaces are available, as well as other means of drying.

Certainly, power and fuel costs are a key factor in evaluating the applicability of this process for a particular

utility. As mentioned previously, the required purity of the

quicklime must also be considered in the decision-making process.

13.3 Innovative Use and Disposal Options

Several alternative uses and disposal methods for residual solids are presented in Slib, Schlamm, Sludge

(Cornwell and Koppers, 1990) and in recent articles by

Copeland et al. (1994, 1995). These alternatives include

options such as land application and reclamation, supplementation of a soil matrix, blending with compost

material, and brick and cement production. Implementation of these alternative methods may prove to be more

cost-effective than traditional disposal methods and will

likely achieve a high degree of public acceptance.

The regulatory framework for acceptance of these options varies among states. It appears that few states

have a well-defined program for approving these uses

as they apply to water treatment plant (WTP) residuals.

Disposal of wastewater treatment plant (WWTP)

biosolids is more clearly regulated than disposal of WTP

residuals, primarily because WWTP biosolids are covered under Part 503 regulations. While Part 503 specifically excludes WTP residuals, many states are applying

similar approaches and standards to the drinking water

industry. As the characteristics of water and wastewater

residuals are very different from one another, more definitive guidance and regulations are needed to address

WTP residuals specifically. Key factors and considerations associated with these innovative use and disposal

alternatives are discussed below.

13.3.1 Beneficial Use

Beneficial use options for WTP residuals include land

application and reclamation, supplementation of a soil

244
matrix, and blending with compost material. Potential

concerns have been raised about aluminum levels in

alum coagulant residuals and their tendency to bind up

available phosphorus in the soil. Phosphorus is needed

as a nutrient for vegetation. To overcome this potential

deficiency, supplemental fertilization may be necessary.

This may be of particular relevance when using residuals for land application, soil matrix supplementation, or

blending with compost material. Some potential benefits

associated with using coagulant solids include improvements in aeration and moisture retention (Copeland et

al., 1994, 1995).

Soil conditioning with lime softening residuals is also a

beneficial use. Farmers often add calcium carbonate to

soils to counteract the chemical reduction that results

from nitrogen application and to maintain an appropriate

soil pH. Any difficulty associated with this use is due

more to farmers’ lack of familiarity with handling the

residual solids than to the residuals themselves.

Blending with compost requires further definition of mix

ratios and maintenance of temperature requirements.

As discussed in Copeland et al. (1994), different mix

ratios must be evaluated and the optimum defined, both

in terms of the existing regulatory framework as well as

the end-use requirements. Maintenance of appropriate

temperature during composting is essential to a successful project.

13.3.2 Disposal Options

Some research and actual implementation has been

conducted in the development of co-products such as

bricks and cement. Again, limited application has occurred in the United States. Willingness of all parties to

conduct preliminary research as well as bench- and

pilot-scale studies is essential.

One successful project involves the Santa Clara Valley

Water District, where bricks are constructed using coagulant residuals. The project took a great deal of effort

and time, however. Careful consideration of associated

costs, benefits, and other economic factors must also be

considered in the development of a co-product.

245
Chapter 14

References

When an NTIS number is cited in a reference, that

document is available from:

 National Technical Information Service

 5285 Port Royal Road

 Springfield, VA 22161

 703-487-4650

 1-800-553-NTIS

Albrecht, A.E. 1972. Disposal of alum sludges. JAWWA 64(1):46-52.

Alfa-Laval Sharples. 1994. Manufacturer’s product literature. AlfaLaval Sharples, 955 Mearns Rd., Warminster, PA 18974.

Ambrose, R.B., Jr., and T.O. Barnwell, Jr. 1989. Environmental software

at the U.S. Environmental Protection Agency’s Center for Exposure

Assessment Modeling. Environmental Software 4(2):76-93.

American Society of Civil Engineers (ASCE)/American Water Works

Association (AWWA). 1990. Water treatment plant design, 2nd ed.

New York, NY: McGraw-Hill.

ASTM. 1982. Standard definitions of terms relating to water. D1129.

Philadelphia, PA: American Society for Testing and Materials.

AWWA. 1986. 1984 water utility operating data. Denver, CO: American Water Works Association.

AWWA/AWWARF. 1992. Water industry database: Utility profiles. Report no. 20293. Denver, CO: American Water Works Association/American Water Works Association Research Foundation.

November.

American Water Works Association Research Foundation (AWWARF). 1969. Disposal of wastes from water treatment plants.

JAWWA 61:541, 619.

AWWA Water Treatment Residuals Management Committee, American Society of Civil Engineers. 1991. State regulatory approaches

for water plant wastes. Proceedings of the 1991 Annual AWWA

Conference and Exposition. Presented at Preconference Seminar,

Current Perspectives on Water Treatment Plant Sludge Disposal:

Practices and Regulations, Philadelphia, PA. Denver, CO: American Water Works Association. June 23.

Andelman, J.B. 1973. Incidence, variability, and controlling factors for

trace elements in natural, fresh waters. In: Singer, P.C., ed. Trace

metals and metal-organic interactions in natural waters. Ann Arbor,

MI: Ann Arbor Science Publishers. pp. 57-88.

Anderson, D.T., and M.J. Hammer. 1973. Effects of alum addition on

activated sludge biota. Water and Sewage Works. January.

Andritz Ruthner, Inc. 1994. Manufacturer’s product literature. Andritz

Ruthner, Inc., 1010 Commercial Blvd. S., Arlington, TX 76017.

Assadian, N.W., and L.B. Fenn. 1995. Beneficial use of combined

biosolids and water plant residuals on agricultural land. In: Proceedings from the AWWA/WEF Joint Residuals Management

Conference, Kansas City, MO. July 23-26.

Baker, J.P. 1982. Effects on fish of metals associated with acidification. In: Proceedings of the International Symposium on Acidic

Precipitation Fishing Impacts in Northeastern N. America. Ithaca,

NY: Cornell University. pp. 165-176.

Baker, J.P., and C.L. Schofield. 1982. Aluminum toxicity to fish in

acidic waters. Water, Air, and Soil Pollution 18:289-309.

Barth, E.F., M.B. Ettinger, B.V. Salotto, and G.N. McDermott. 1965.

Summary report on the effects of heavy metals on the biological

treatment processes. J. Water Pollut. Control Fed. 37(1):86-96.

Bassett, R.L., and D.C. Melchoir. 1990. Chemical modeling of aqueous systems. In: Melchoir, D.C., and R.L. Bassett, eds. Chemical

modeling in aqueous systems II. American Chemical Society Symposium Series 416. Washington, DC. pp. 1-14.

Bennett, D.L. 1978. The efficiency of water treatment processes in

radium removal. JAWWA 70(12):698-701.

Birchall, J.D., C. Exley, J.S. Chappel, and M.J. Phillips. 1989. Acute

toxicity of aluminum to fish eliminated in silicon-rich acid waters.

Nature 338(6211):146-148.

Bowen, P.T., M.K. Jackson, and R.A. Corbitt. 1991. Sludge treatment,

usage, and disposal. J. Water Pollut. Control Fed. 63(4):406-414.

Bratby, J., C. Baugh, and W. Ambrose. 1993. Water plant sludge

dewatering by recessed chamber filter press. In: Proceedings of

the AWWA/WEF Joint Residuals Management Conference, Phoenix, AZ. December 5-8.

Brink, W.L., R.J. Schliekelman, D.L. Bennett, C.R. Bell, and I.M. Markwood. 1978. Radium removal efficiencies in water treatment processes. JAWWA 70(1):31-35.

Brinkhurst, R.O., A.L. Hamilton, and H.B. Herrington. 1968. Components of the bottom fauna of the St. Lawrence, Great Lakes. No.

PR 33. Great Lakes Institute, University of Toronto, Toronto, Ontario, Canada.

Brobst, R. 1994. Survey of 12 land-applied WTP residuals. Unpublished data. U.S. Environmental Protection Agency Region 8.

Denver, CO.

Brown, D.J.A. 1983. Effect of calcium and aluminum concentrations

on the survival of Brown Trout (Salmo trutta) at low pH. Bull. Env.

Contam. Toxicol. 30:582-587.

Brown and Caldwell, Inc. 1990. City of Phoenix water residuals management study. Study produced for the City of Phoenix, AZ.

Brusewitz, S. 1984. Aluminum. Stockholm, Sweden: University of

Stockholm, Institute of Theoretical Physics. 203:138.

246
Buckler, D.R., P.M. Mehrle, L. Cleveland, and F.J. Dwyer. 1987. Influence of pH on the toxicity of aluminum and other inorganic contaminants to the East Coast striped bass. Water, Air, and Soil

Pollution. 35:97-106.

Bugbee, G.J., and C.R. Frink. 1985. Alum sludge as a soil amendment: Effects on soil properties and plant growth. Connecticut

Agricultural Experiment Station Bulletin 827.

Burrows, W.D. 1977. Aquatic aluminum: Chemistry, toxicology, and

environmental prevalence. CRC Critical Reviews in Environmental

Control. Boca Raton, FL: CRC Press. pp. 167-216.

Cairns, J., Jr. 1977. Quantification of biological integrity. In: Ballentine,

R.K., and L.J. Guarraia, eds. The integrity of water. U.S. Environmental Protection Agency, Office of Water and Hazardous Materials, Washington, DC. U.S. Government Printing Office Stop No.

055-001-01068-1.

Cairns, J., Jr., J.S. Crossman, K.L. Dickson, and E.E. Herricks. 1971.

The recovery of damaged streams. ASB Bulletin 18(3):79-106.

Calkins, R.J., and J.T. Novak. 1973. Characteristics of chemical

sludges. JAWWA 65(6):523.

Campbell, P.G.C., M. Bisson, R. Bougie, A. Tessier, and J. Villeneuve.

1983. Speciation of aluminum in acidic freshwaters. Anal. Chem.

55:2246-2252.

CDM. 1989. City of Norfolk, 37th Street Water Treatment Plant,

sludge disposal study. Raleigh, NC: Camp Dresser & McKee.

January.

Chang, A.C., A.L. Page, and J.E. Warneke. 1983. Soil condition effects of municipal sludge compost. J. Environ. Eng. 109:574.

Che, M.D., T.J. Logan, S.J. Traina, and J.M. Bigham. 1988. Properties

of water treatment sludges and their effectiveness as agricultural

limestone substitutes. Columbus, OH: Ohio State Agricultural Research and Development Center. 60:674-680.

Chisholm, J.L., and S.C. Downs. 1978. Stress and recovery of aquatic

organisms as related to highway construction along Turtle Creek,

Boone County, West Virginia. U.S. Geological Survey Water Supply Paper 2055. Washington, DC.

Clifford, D.A. 1990. Removal of radium from drinking water. In: Cothern, C.R., and P. Rebers, eds. Radon, radium, and uranium in

drinking water. Chelsea, MI: Lewis Publishers. pp. 225-247.

Clifford, D.A., W. Vijjesworapu, and S. Subramonian. 1988. Evaluating

various adsorbents and membranes for removing radium from

groundwater. JAWWA 80(7):94.

Cobb County-Marietta Water Authority. 1994. Data from the Cobb

County-Marietta Water Authority, 1660 Barnes Mill Road, Marietta,

GA 30062.

Colorado Department of Health. 1986. Domestic sewage sludge regulations, 5CCR 1003-7.

Colorado Department of Health. 1990. Colorado regulation pertaining

to solid waste disposal sites and facilities promulgated under the

Solid Waste Disposal Sites and Facilities Act. Title 30, Part 1, CRS

as amended.

Colorado Springs Utilities. 1994. Data from the Colorado Springs

Utilities, Water Resources Department, P.O. Box 1103, Colorado

Springs, CO 80947.

Copeland, A.M., C. Vandermeyden, and D.A. Cornwell. 1994. Beneficial use alternatives for water treatment plant residuals. National

Conference on Environmental Engineering, Boulder, CO.

Copeland, A.M., C. Vandermeyden, and D.A. Cornwell. 1995. Residuals benefits. Civil Engineering Journal. January: 70-72.

Cornwell, D., M.M. Bishop, R.G. Gould, and C. Vandermeyden. 1987.

Handbook of practice: Water treatment plant waste management.

Denver, CO: American Water Works Association Research

Foundation.

Cornwell, D.A., and H.M.M. Koppers, eds. 1990. Slib, schlamm,

sludge. Denver, CO: American Water Works Association Research

Foundation and KIWA Ltd.

Cornwell, D.A., and R. Lee. 1993. Recycle stream effects on water

treatment. Denver, CO: American Water Works Association Research Foundation.

Cornwell, D.A., and R.G. Lee. 1994. Waste stream recycling: Its effect

on water quality. JAWWA 86(11):50-63.

Cornwell, D.A., C. Vandermeyden, G. Dillow, and M. Wang. 1992.

Landfilling of water treatment plant coagulant sludges. Denver,

CO: American Water Works Association Research Foundation/American Water Works Association.

Dentel, S., M. Abu-Orf, and N. Griskowitz. 1993. Guidance manual

for polymer selection in wastewater treatment plants. Water Environment Research Federation.

Dempsey, B.A., and H.A. Elliott. 1990. Land application of water

treatment sludges. In: Proceedings of the 44th Purdue Industrial

Waste Conference, Purdue University. W. Lafayette, IN: Lewis

Publishers.

Dixon, K.L. 1993. Hazardous waste classification. In: Proceedings of

the AWWA/WEF 1993 Joint Residuals Management Conference.

Phoenix, AZ. December 5-8.

Dixon, K.L., R.G. Lee, and R.H. Moser. 1988. Water treatment plant

residuals: A management strategy for the Pennsylvania region.

Voorhees, NJ: American Water Works Service Company. June 23.

Donovan, J.F., and W.E. Toffey. 1993. Impact of industrial pretreatment programs on biosolids quality in Boston and Philadelphia.

In: Proceedings of the AWWA/WEF 1993 Joint Residuals Management Conference, Phoenix, AZ. December 5-8.

DPRA, Inc. 1992. Telephone survey of municipal sewer rates, September 15-17. DPRA, Inc., 332 Minnesota St., St. Paul, MN 55101.

Dragun, J. 1988. The soil chemistry of hazardous materials. p. 77.

Driscoll, C.T., Jr. 1985. Aluminum in acidic surface waters: Chemistry,

transport, and effects. Environ. Health Perspectives 63:93-104.

Driscoll, C.T., J.P. Baker, J.J. Bisogni, and C.L. Schofield. 1980. Effects of aluminum speciation on fish in dilute acidified waters.

Nature 284:161-164.

Driscoll, C.T., Jr., and W.D. Schecher. 1988. Aluminum in the environment. In: Sigel, H., ed. Metal ions in biological systems: Aluminum and its role in biology, Vol. 24. New York, NY: Decker

Publishing. pp. 1-41.

Driscoll, C.T., Jr., B.J. Wyskowski, and P. Destaffan. 1987. Chemistry

and transfer of aluminum in a forested watershed in the Adirondack region of New York, USA. Presented at the 194th National

Meeting of the American Chemical Society, New Orleans, LA.

Dulin, D.E., and W.R. Knocke. 1989. The impact of incorporated

organic matter on the dewatering characteristics of aluminum hydroxide sludges. JAWWA 81(5):74.

EET, Inc. 1992. Residuals Management Study for West Virginia

American Water Company. Submitted to the American Water

Works Service Company, Inc. Newport News, VA: Environmental

Engineering & Technology, Inc. November.

EIFAC. 1969. Water quality criteria for European freshwater fish.

European Inland Fisheries Committee.

247
Elliot, H.A., B.A. Dempsey, and J. DeWolfe. 1988. Land disposal of

drinking water sludge. Presented at 1988 International Winter

Meeting of the American Society of Agricultural Engineers, December 13-16.

Elliott, H.A., B.A. Dempsey, and P.J. Maille. 1990a. Content and fractation of heavy metals in water treatment sludges. J. Environ.

Qual. 19:330-334.

Elliott, H.A., and B.A. Dempsey. 1990b. Land application of water

treatment sludges: Impact and management. Final report. Denver,

CO: American Water Works Association Research Foundation.

Elliott, H.A., and L.M. Singer. 1988. Effect of water treatment sludge

on growth and element composition of tomato (Lycopersicon esculentum) shoots. Communications in Soil Science and Plant

Analysis 11:969-980.

Emig, D.K. 1979. Excessive alum causes indigestion. Water and

Waste Engineering. April.

Esernick, E.L., J.L. Maas-Diepeveen, and C.J. Leeuwen. 1991. Combined effects of metals—an ecotoxicological evaluation. Water

Res. 25(6):679-688.

Evans, R.L., D.H. Schnepper, and T.H. Hill. 1979. Impact of wastes

from a water treatment plant: Evaluation procedures and results.

Illinois State Water Survey Circular 135. p. 39.

Evans, R.L., D.H. Schnepper, T.E. Hill, and D.L. Hullinger. 1982.

Waste from the water treatment plant at Alton and its impact on

the Mississippi River. Illinois State Water Survey Circular 156. p.

62.

Evans, R.E., S.B. Brown, and T.J. Hara. 1988. The effects of aluminum and acid on the gill morphology in rainbow trout, Salmo

gardneri. Environ. Biol. Fishes 22(4):299-311.

Foley, J.F., K. Folks, and K.J. Yakich. 1989. Treating water filter plant

sludge at a wastewater treatment plant. Presented at the 1989

AWWA/WPCF Joint Residuals Management Conference, San

Diego, CA. August 13-16.

Foster, W.S. 1975. Get the water out of alum sludge. American City

and County 90(9).

Foth, H.D., and B.G. Ellis. 1988. Soil fertility. New York, NY: Wiley &

Sons.

Freeman, R.A., and W.H. Everhart. 1971. Toxicity of aluminum hydroxide complexes in neutral and basic media to rainbow trout.

Trans. Am. Fish. Soc. 100:644.

Freeman, R.A., and W.H. Everhart. 1973. Recovery of rainbow trout

from aluminum poisoning. Trans. Am. Fish Soc. 100:152.

Funk, W.H., H.R. Gibbons, and S.K. Bhagat. 1977. Nutrient interaction by large scale aluminum sulfate treatment. Presented at Conference on Mechanics of Lake Restoration, University of

Wisconsin, Madison, WI, April 25-28.

Gammon, J.R. 1970. The effect of inorganic sediment on stream

biota. Water Poll. Cont. Res. Series. 18050 DWC 12/70. Washington, DC.

Gandhi, R.L. 1992. Slurry and sludge piping. In: Piping handbook,

6th ed. New York, NY: McGraw-Hill.

Geertsema, W.S., W.R. Knocke, J.T. Novak, and D. Dove. 1994.

Long-term effects of sludge application to land. JAWWA 86(11):64-

74.

George, D.B., S.G. Berk, V.D. Adams, E.L. Morgan, R.O. Roberts,

C.A. Holloway, R. Lott, L.K. Holt, R.S. Ting, and A.W. Welch. 1991.

Alum sludge in the aquatic environment. Denver, CO: American

Water Works Association Research Foundation.

George, D.B., S.G. Berk, V.D. Adams, R.S. Ting, R.O. Roberts, L.H.

Parks, and R.C. Lott. 1995. Toxicity of alum sludge extracts to

a freshwater alga, protozoan, fish, and marine bacterium. In:

Archives of Environmental Contamination and Toxicology 29(2):

149-158.

Goodnight, C.J. 1973. The use of aquatic macroinvertebrates as indicators of aquatic pollution. Transactions of the American Microscopical Society 92(1):1-13.

Goss, G.G., and C.M. Wood. 1988. The effects of acid and acid-aluminum exposure on circulating plasma cortisol levels and other

blood parameters in rainbow trout, Salmo gairdneri. J. Fish Biol.

32(1):63-76.

Grabarek, R.J., and E.C. Krug. 1987. Silviculture application of alum

sludge. JAWWA 79(6):84-88.

Grabarek, R.J., and E.C. Krug. 1989. Silviculture application of alum

sludge. In: Sludge: Handling and disposal. Denver, CO: American

Water Works Association. pp. 97-101.

Great Lakes Upper Mississippi River Board of State Public Health

and Environmental Managers. 1992. Recommended standards for

water works. New York, NY: Health Education Services.

Hahn, N.A., Jr. 1989. Disposal of radium removal from dirty water. In:

Sludge: Handling and disposal. Denver, CO: American Water

Works Association.

Hall, R.J., C.T. Driscoll, G.E. Likens, and J.M. Pratt. 1985. Physical,

chemical and biological consequences of episodic aluminum additions to a stream. Limnology and Oceanography 30:212-218.

Hall, W.S., and L.W. Hall, Jr. 1989. Toxicity of alum sludge to Ceriodaphnia dubia and Pimephales promplas. Bull. Env. Contam. Toxicol. 42(5):791-798.

Harberg, B., and P. Heppler. 1993. Co-application of water treatment

plant residuals and wastewater plant biosolids to dry land winter

wheat in Colorado. In: Proceedings of the AWWA/WEF Joint Residuals Management Conference, Phoenix, AZ. December 5-8.

Havas, M. 1985. Aluminum bioaccumulation and toxicity to Daphnia

magna in soft water at low pH. Can. J. Fish. Aquat. Sci. 42:1741-1748.

Havas, M., and T.C. Hutchinson. 1983. Effects of low pH on the

chemical composition of aquatic invertebrates from tundra ponds

at Smoking Hills, N.W.T., Canada. Can. J. Zool. 61:241-249.

HDR Engineering, Inc. 1995. Study to develop BCT for NPDES permit

limits for water treatment plants. Draft report to the City of Phoenix,

Water Services Department, Phoenix, AZ.

Heil, D.M. 1988. The influence of water treatment sludges on the

growth of sorghum-sudangrass. M.S. thesis. Colorado State University, Fort Collins, CO.

Heil, D.M., and K.A. Barbarick. 1989. Water treatment sludge influence on the growth of sorghum-sudangrass. J. Environ. Qual.

18:292-298.

Holmgren, G.G.S., M.W. Myer, R.H. Chaney, and R.B. Daniels. 1993.

Cadmium, lead, zinc, copper, and nickel in agricultural soils of the

United States of America. J. Environ. Qual. 22:335-348.

HPD, Inc. 1994. Product literature. Naperville, IL.

Hsu, C.Y., and W.O. Pipes. 1973. Aluminum hydroxide effects on

wastewater treatment processes. JWPCF 45(4):681-697.

Hundt, T.R. 1986. Aluminum-fulvic acid interactions: Mechanisms and

applications. In: Proceedings of the Annual AWWA Conference,

Denver, CO. June 22-26. pp. 1199-1201.

248
Hunn, J.B., L. Cleveland, and E.E. Little. 1987. Influence of pH and

aluminum on developing brook trout in low calcium water. Environmental Pollution 43:63-73.

Illinois Department of Energy and Natural Resources. 1987. Wastes

from water treatment plants: Literature review, results of an Illinois

survey, and effects of alum sludge application to cropland.

Peoria, IL.

Illinois Department of Nuclear Safety. 1984. Management of water

treatment sludge containing elevated levels of radium.

Infilco Degremont. 1994. Manufacturer’s product literature. Infilco Degremont, 2924 Emergy Wood Parkway, Richmond, VA 23294.

Ingram, W.T. 1969. Glossary: Water and wastewater control engineering. APHA, American Society of Civil Engineers, American Water

Works Association, Water Pollution Control Federation. p. 40.

Jelinek, R.T., and T.J. Sorg. 1988. Operating a small full-scale ion

exchange system for uranium removal. JAWWA 80(8):79-83.

Public Works. 1992. Job postings. July.

John Carollo Engineers, Inc. 1994. Unpublished material. John

Carollo Engineers, Inc., 3877 N. Seventh St., Suite 400, Phoenix,

AZ 85014.

Karlson-Norrgren, L., I. Borklund, O. Lujungberg, and P. Runn. 1986.

Acid waters and aluminum exposure: Experimentally induced gill

lesions in brown trout, Salmo trutta. J. Fish Diseases 9:11-25.

Kaverkamp, J.F., W.A. MacDonald, W.R. Lillie, and A. Lutz. 1983.

Joint toxicity of mercury and selenium in Salmonid eggs. Arch.

Environ. Contam. Toxicol. 12(4):415-419.

Kaviraj, A., and S.K. Konar. 1983. Lethal effects of mixtures of mercury, chromium, and cadmium on fish. Environ. Ecol. 1(3):

153-158.

Kawczyinski, E., and V. Achterrman. 1991. A water industry data base

report. In: Proceedings of the AWWA/WEF Joint Residuals Management Conference, Research Triangle Park, NC.

Kay, J. 1986. Cadmium accumulation and protein binding patterns in

tissues of the rainbow trout, Salmo gairdneri. Conference on HighAffinity Metal-Binding Proteins in Non-mammalian Species. Environ. Health Perspect. 65:133-139.

King, C.H., C.P. Houck, and L.L. Deadmond. 1988. Land application

of Kodak Colorado Division Wastewater Treatment Plant sludge.

Presented at Residuals Management Specialty Conf. Water Pollution Control Federation, Athens, GA. April 19-20.

Knocke, W.R., J.R. Harmon, and B.E. Dulin. 1987. Effects of coagulation on sludge thickening and dewatering. JAWWA 79:89. 

Knocke, W.R., M.M. Ghosh, and J.T. Novak. 1980. Vacuum filtration

of metal hydroxide sludges. J. Environ. Eng. Division, ASCE.

106(2):363.

Knocke, W.R., C.M. Dishman, and G.F. Miller. 1993. Measurement of

chemical sludge floc density and implications related to sludge

dewatering. Water Environ. Res. 65:735.

Koorse, S.J. 1993a. The role of residuals disposal laws in treatment

plant design. JAWWA 85(10):57.

Koorse, S.J. 1993b. Toxic regulations take hold. Water Environment

and Technology. January.

Koplish, D.E., and M.E. Watson. 1991. Demonstration and feasibility

of combined dewatering of alum and waste sludges and the economic benefits of a single site facility. Presented at the 1991

AWWA/WPCF Joint Residuals Management Conference, Durham, NC. August 11-14.

Koplish, D.E., J. McMahon, and J. Valek. 1995. Efficiencies of centralized residuals dewatering and co-disposal. In: Proceedings of

the AWWA/WEF Joint Residuals Management Conference, Kansas City, MO. July 23-26.

Lagvankar, A.L., and R.S. Gemmell. 1968. A size-density relationship

for flocs. JAWWA 60(9):1040.

Lamb, D.S., and G. Bailey. 1981. Acute and chronic effects of alum

to Midge larva. Bulletin Env. Contam. Toxicol. 27:59.

Lapedes, D.N. 1978. Dictionary of scientific and technical terms, 2nd

ed. New York, NY: McGraw-Hill.

Lassovszky, P., and S. Hathaway. 1983. Treatment technologies to

remove radionuclides from drinking water. Preconference report

for the National Workshop on Radioactivity in Drinking Water.

Washington, DC: U.S. Environmental Protection Agency.

League of Minnesota Cities. 1991. A research memo for city officials:

Minnesota water and sewer rates. League of Minnesota Cities,

183 University Ave., E., St. Paul, MN 55101.

Lee, S.Y., and E.A. Bondietti. 1983. Removing uranium from drinking

water by metal hydroides and anion-exchange resin. JAWWA

75(10):536-540.

Lee, R.G., H. Eckhardt, N. Wortel, and H.M.M. Koppers. 1990. Toxic

concerns regarding water plant wastes. In: Cornwell, D., and

H.M.M. Koppers, eds. Slib, schlamm, sludge. Denver, CO: American Water Works Association Research Foundation.

Lin, S.M. 1988. Effects of alum sludge on corn and soybeans. In:

Proceedings of the Joint CSCE-ASCE National Conference on

Environmental Engineering, Specialty Conference. July 13-15.

Lin, S.D. 1989. Impacts of alum sludge discharge from a mid-sized

water treatment plant on a small stream. Presented at the 1989

AWWA/WPCF Joint Residuals Management Conference, San Diego, CA. August 13-16.

Lin, S., and C. Green. 1987. Wastes from water treatment plants:

Literature review, results of an Illinois survey, and effects of alum

sludge application to cropland. Illinois State Water Survey.

November.

Logsdon, G.L., ed. 1991. Slow sand filtration. Task Committee Report.

New York, NY: American Society of Civil Engineers.

Lowry, J.D., and J.E. Brandow. 1985. Removal of radon from water

supplies. J. Environ. Eng. 3(4):511-527.

Mackenthun, K.M. 1973. Toward a cleaner aquatic environment.

Washington, DC: U.S. Government Printing Office.

Malate, H. 1986. Effects of aluminum in hard, acid water on metabolic

rate, blood gas tensions and sonic status in rainbow trout. J. Fish

Biol. 29:187-198.

Malmrose, P.E., and T.A. Wolfe. 1994. Recent advances in technologies for dewatering coagulant residuals. AWWA Conference on

Development of Programs for Engineering Tomorrow’s Water Systems, Cincinnati, OH.

Mance, G. 1987. Pollution threat of heavy metals in aquatic environments. New York, NY: Elsevier Applied Science.

Martell, A.E., and R.J. Motekaitus. 1989. Coordination chemistry and

speciation of Al(III) in aqueous solution. In: Lewis, T.E., ed. Environmental Chemistry and Toxicology of Alum. Chelsea, MI: Lewis

Publishers.

McDonald, D.G., R.L. Walker, and P.R.H. Wilkes. 1983. The interaction of environmental calcium and low pH on the physiology of the

rainbow trout, Salmo gairdneri. J. Eper. Biol. 102:141-155.

249
McNatt, R.M. 1977. Zooplankton species composition, abundance,

and distribution in a desert, pumped-storage reservoir. Ph.D. Dissertation. Arizona State University, Tempe, AZ.

McTigue, N., and D.A. Cornwell. 1994. The hazardous potential of

activated carbons used in water treatment. Denver, CO: American

Water Works Association Research Foundation.

McTigue, N.E., D.A. Cornwell, and A.T. Rolan. 1989. Impact of water

plant waste discharge on wastewater plants. Presented at the

1989 AWWA/WPCF Joint Residuals Management Conference,

San Diego, CA. August 13-16.

Mesa WTP. 1994. Mesa Water Treatment Plant, Colorado Springs

Utilities, Water Resources Department. P.O. Box 1103, Colorado

Springs, CO 80947.

Metcalf & Eddy. 1991. Wastewater engineering: Treatment, disposal,

and reuse, 3rd ed. New York, NY: McGraw-Hill.

Meyers, A.G., V.L. Snoeyink, and D.W. Snyder. 1985. Removing barium and radium through calcium cation exchange. JAWWA

77(5):60-66.

Mickley, M., R. Hamilton, L. Gallegos, and J. Truesdall. 1993.

Membrane concentrate disposal. Denver, CO: American Water

Works Association/American Water Works Association Research

Foundation.

Missouri Department of Natural Resources. 1985. Agricultural use of

municipal wastewater sludge—a planning guide. Division of Environmental Quality.

Mulbarger, M.C., S.R. Copas, J.R. Kordic, and F.M. Cosh. 1981.

Pipeline friction losses for wastewater sludges. J. Water Pollut.

Control Fed., Vol. 53. (August).

Mukhopadhayay, M.K., and S.K. Konar. 1985. Effects of copper, zinc,

and iron mixture on fish and aquatic ecosystem. Environ. Ecol.

3(1):58-64.

Muniz, I.P., and H. Leivestad. 1980. Toxic effects of aluminum on the

brown trout, Salmo trutta. In: Drablos, D., and A. Tollam, eds.

Ecological impact on acid precipitation. Mysen, Norway: Grefslie

Press.

MWRDGC. 1994. Metropolitan Water Reclamation District of Greater

Chicago, 100 East Erie St., Chicago, IL 60611.

New Jersey DEPE. 1993. New Jersey statewide sludge management

plan. New Jersey Department of Environmental Protection and

Energy. November.

Novak, J.T. 1986. Historical and technical perspective of sludge treatment and disposal. Presented at Recent Advances in Sludge

Treatment and Disposal, American Water Works Association. Preconference Seminar Proceedings, Washington, DC. June.

Novak, J.T. 1989. Historical and technical perspective at sludge treatment and disposal. In: Sludge: Handling and disposal. Denver,

CO: American Water Works Association.

Novak, J.T., and D.C. Calkins. 1973. Sludge dewatering and its physical properties. JAWWA. 67(18).

O’Connor, J.T., and J.T. Novak. 1978. Management of water treatment

plant residues. In: Proceedings of the AWWA Water Treatment

Waste Disposal Seminar, Atlantic City, NJ. AWWA.

Olsen, R.D. 1975. The phytoplankton ecology of a desert lake. Ph.D.

Dissertation. Arizona State University, Tempe, AZ.

Page, A., R.H. Miller, and D.R. Keeney, eds. 1982. Methods of soil

sampling and analysis. Madison, WI: American Society of Agronomy/Soil Science Society of America.

Parrotta, M.J. 1991. Radioactivity in water treatment wastes: A U.S.

Environmental Protection Agency perspective. JAWWA 83(4):134-

140.

Parsons, J.M., and W.L. Waldrip. 1995. A land application study utilizing both lime treated alum sludge from a water treatment plant

and aerobically digested biosolids. In: Proceedings of the AWWA/

WEF Joint Residuals Management Conference, Kansas City, MO.

July 23-26.

Peck, B., J. Russell, R. Ardizzone, and J. Hayes. 1993. A case history

at the Val Vista Water Treatment Plant: Implementing a used water

recovery system. Proceedings of the 1993 AWWA/WEF Joint Residuals Conference.

Pennak, R.W. 1989. Freshwater invertebrates of the United States:

Protozoa to mollusca, 3rd ed. New York, NY: John Wiley & Sons.

Pine Valley WTP. 1994. Data from Pine Valley Water Treatment Plant,

Colorado Springs Utilities, Water Resources Department. P.O. Box

1103, Colorado Springs, CO 80947.

Pontius, F.W., ed. 1990. Water quality and treatment: A handbook of

community water supplies, 4th ed. New York, NY: McGraw-Hill.

Price, S.D., D.M. Owen, G.C. Cline, and K.E. Spiker. 1989. Water treatment plant residuals management for San Diego: The feasibility of

co-disposal. In: Proceedings of the 1989 AWWA/WPCF Joint Residuals Management Conference, San Diego, CA. August 13-16.

Raghu, D., H.N. Hsieh, T. Neilman, and C.T. Yih. 1987. Water treatment

plant sludge as landfill liner. Geotechnical Practice for Waste Disposal 1987, Geotechnical Special Publication 13. New York, NY.

Ramamoorthy, S. 1988. Effects of pH on speciation and toxicity of

aluminum to rainbow trout (Salmo gardneri). Can. J. Fishery

Aquatic Sci. 45:634-642.

Reed, R.C. 1975. The effects of alum sludge on anaerobic digestion.

Master’s thesis. Auburn University, Auburn University, AL.

Reh, C.W. 1978. Proceedings of the Twentieth Annual Public Water

Supply Engineer’s Conference—New Directions for Water Supply

Design and Operation. Department of Civil Engineering, University

of Illinois of Champaign-Urbana. American Water Works Association, Illinois Environmental Protection Agency, and Illinois Section

of AWWA. Champaign, IL. April 3-5.

Reid, G.W., R.Y. Nelson, C. Hall, U. Bonilla, and B. Reid. 1968. Effects

of metallic ions on biological waste treatment processes. Water

and Sewage Works. July.

Rindt, B.A. 1973. The effects of alum sludge on primary sedimentation

and anaerobic digestion. M.S. thesis. Auburn University, Auburn

University, AL.

Roberts, M.H., and R.J. Diaz. 1985. Assessing the effects of alum

sludge discharges into coastal water. In: Proceedings of the

AWWA Seminar on Recent Advances in Sludge Treatment and

Disposal, Washington, DC.

Robertson, R.F., and Y.T. Lin. 1981. Proceedings of the Twenty-Third

Annual Public Water Supply Engineer’s Conference—New Directions for Water Supply Design and Operation. Department of Civil

Engineering, University of Illinois of Champaign-Urbana. American

Water Works Association, Illinois Environmental Protection

Agency, and Illinois Section of AWWA. Champaign, IL. April 21-23.

Robinson, M.P., and J.B. Witko. 1991. Overview of issues and current

state-of-the-art water treatment plant waste management programs. 1991 Annual Conference Proceedings. American Water

Works Association Water Quality for the New Decade, Philadelphia, PA. June 23-27.

250
Rolan, A.T., and J.C. Brown. 1973. Effect of water treatment plant

sludge on the treatment of municipal wastewater. In: Proceedings

of the 53rd Annual Meeting of the North Carolina Section of the

American Waste Water Association and WPCA.

R.S. Means Co., Inc. 1991. 1992 Means building construction cost

data, 50th ed. Kingston, MA: Construction Consultants and

Publishers.

Rueter, J.G., Jr., K.T. O’Reilly, and R.R. Petersen. 1987. Indirect aluminum toxicity to the green alga Scenedesmus through increased

cupricion activity. Environ. Sci. Technol. 21(5):435-438.

Saltonstall, C.W., Jr., and R.W. Lawrence. 1982. Calculation of the

expected performance of reverse osmosis plants. Desalination

42:247-253.

Sanville, W.D., A.R. Gahler, J.A. Searcy, and C.F. Powers. 1976.

Studies on lake restoration by phosphorus inactivation.

EPA/600/3-76/041. U.S. Environmental Protection Agency, Corvallis Environmental Research Laboratory, Corvallis, OR.

Saunders, M., and M. Roeder. 1991. Coagulant review: A critical

assessment. Denver, CO: American Water Works Association Research Foundation.

Schofield, C., and J.R. Trojnar. 1980. Aluminum toxicity to brook trout

(Salvelinus fontinalis) in acidified waters. In: Toribara, T.Y., M.W.

Miller, and P.E. Morrow, eds. Polluted Rain. New York, NY: Plenum

Press. pp. 341-363.

Shacklette, H.T., and J.G. Boerngen. 1984. Element concentrations

in soils and other surficial materials of the conterminous United

States. U.S. Geological Survey Professional Paper 1270.

Smith, R.W., and J.D. Hem. 1972. Effect of aging on aluminum hydroxide complexes in dilute aqueous solutions. Chemistry of aluminum in natural water. U.S. Geological Survey Water Supply

Paper 1827-D. Washington, DC: U.S. Government Printing Office.

Snoeyink, V.L., C.K. Jongeward, A.G. Myers, and S.K. Ritcher. 1989.

Characteristics and handling of wastes from groundwater treatment systems. In: Sludge: Handling and disposal. Denver, CO:

American Water Works Association.

Sorensen, E.M. 1991. Metal poisoning in fish. Boca Raton, FL: CRC

Press.

Sorg, T.J., R.W. Forbes, and D.S. Chambers. 1980. Removal of radium-226 from Sarasota County, FL, drinking water by reverse

osmosis. JAWWA 72(4):230-237.

Sorg, T.J. 1988. Methods for removing uranium from drinking water.

JAWWA 80(7):105-111.

Sorg, T.J., and G.S. Logsdon. 1980. Treatment technology to meet

the interim primary drinking water regulations for inorganics, Part

5. JAWWA 72(7):411.

Sorg, T.J., D.A. Lytle, J. Egan, and Z. Zhang. 1991. Characterization

of wastes (brine) from uranium removal plant. Presented at

AWWA/WEF Conference on Future of Residuals Management,

Research Triangle Park, NC. August 11-19.

Spehar, R.L., and J.T. Fiandt. 1986. Acute and chronic effects of

water-quality criteria-based metal mixtures on three aquatic species. Environ. Toxicol. Chem. 5(10):917-931.

Suffern, J.S., C.M. Fitzergerald, and A.T. Szluha. 1981. Trace metal

concentrations in oxidation ponds. J. Water Pollut. Control Fed.

53(11):1599-1608.

Tanner, S.A., and J.E. Ongerth. 1990. Evaluating the performance of

slow sand filters in northern Idaho. JAWWA 82(12):51.

Tebo, L.B. 1965. Effects of siltation, resulting from improper logging,

on the bottom fauna of a small trout stream in the southern Appalachians. Progressive Fish Culturist 17:64.

Thomann, R.V., and J.A. Mueller. 1987. Principles of surface water

quality modeling and control. New York, NY: Harper and Row

Publishers. p. 644.

Tsang, K.R., and R.L. Hurdle. 1991. The co-disposal of water and

wastewater treatment residuals—feasibility study. Presented at

the 1991 AWWA/WPCF Joint Residuals Management Conference, Durham, NC. August 11-14.

U.S. Department of Health and Human Services (DHHS). 1990. Toxicological profile for aluminum. Washington, DC.

U.S. EPA. 1976a. Determination of radium removal efficiencies in

Illinois water supply treatment processes. Technical Note

ORP/TAD-76-2. NTIS PB296153. Washington, DC.

U.S. EPA. 1976b. Determination of radium removal efficiencies in

Iowa water supply treatment processes. Technical Note ORP/TAD76-1. NTIS PB259509. Washington, DC.

U.S. EPA. 1977. Manual of treatment techniques for meeting the

interim primary drinking water regulations. EPA/600/8-77/995. Cincinnati, OH.

U.S. EPA. 1978a. Beneficial disposal of water purification plant sludges

in wastewater treatment. EPA/600/2-78/089. Cincinnati, OH.

U.S. EPA. 1978b. Process design manual: Municipal sludge landfills.

EPA/625/1-78/010. p. 5-9.

U.S. EPA. 1979a. Estimating water treatment plant costs. EPA/600/2-

79/162a. Cincinnati, OH.

U.S. EPA. 1979b. Process design manual for sludge treatment and

disposal. EPA/625/1-79/011. Cincinnati, OH.

U.S. EPA. 1983. Methods for chemical analysis of water and wastes.

Revised ed. EPA/600/4-79/020. Cincinnati, OH.

U.S. EPA. 1984a. Barium and radium in water treatment plant wastes.

EPA/600/2-85/006. Cincinnati, OH.

U.S. EPA. 1984b. Environmental regulations and technology: Use and

disposal of municipal wastewater sludge. EPA/625/10-84/003. As

printed in Metcalf and Eddy, Inc., Wastewater Engineering,

McGraw-Hill. p. 772.

U.S. EPA. 1984c. MINTEQ—a computer program for calculating

aqueous geochemical equilibria. EPA/600/3-84/032. Washington,

DC.

U.S. EPA. 1985a. Slow sand filter maintenance costs and effects of

water quality. EPA/600/2-85/056. NTIS PB85199669. Cincinnati,

OH.

U.S. EPA. 1985b. Water quality assessment: A screening procedure

for toxic and conventional pollutants in surface and groundwater.

EPA/600/6-85/002b. NTIS PB86122504.

U.S. EPA. 1986a. National water quality inventory. 1986 Report to

Congress. EPA/440/4-87/008. Washington, DC.

U.S. EPA. 1986b. Test methods for evaluating solid waste: Physical/Chemical methods, 3rd ed. SW-846. NTIS PB88239223.

Washington, DC.

U.S. EPA. 1986c. Water quality criteria 1986. EPA/440/5-86/001.

NTIS PB87226759. Washington, DC.

U.S. EPA. 1987. Removal of uranium from drinking water by ion

exchange and chemical clarification. EPA/600/52-87/076. Cincinnati, OH.

251
U.S. EPA. 1988a. Ambient water quality criteria for aluminum.

EPA/440/5-86/008. Washington, DC.

U.S. EPA. 1988b. Radium removal for a small community water supply system. EPA/600/52-88/039. Cincinnati, OH.

U.S. EPA. 1988c. Reverse osmosis treatment to remove inorganic contaminants from drinking water. EPA/600/52-87/109. Cincinnati, OH.

U.S. EPA. 1989. POTW sludge sampling and analysis guidance document. U.S. Environmental Protection Agency Permits Division.

NTIS PB93227957. August.

U.S. EPA. 1990a. National sewage sludge survey: Availability of information and data, and anticipated impacts on proposed regulations. Proposed rule. Fed. Reg. 55:47210. November 9.

U.S. EPA. 1990b. Radon removal techniques for small community

public water supplies. EPA/600/52-90/036. Cincinnati, OH.

U.S. EPA. 1990c. Suggested guidelines for the disposal of drinking

water treatment wastes containing naturally occurring radionuclides. EPA/570/F-90/018. Washington, DC.

U.S. EPA. 1991a. MINTEQA2/PRODEFA2, a geochemical assessment model for environmental systems: Version 3.0 user’s manual. EPA/600/3-91/021. Washington, DC.

U.S. EPA. 1991b. National primary drinking water regulations; radionuclide; proposed rule. Fed. Reg. 56(138):33,050. July 18.

U.S. EPA. 1992a. Organic air emissions from waste management facilities. Seminar Publication. EPA/625/R-92/003. Washington, DC.

U.S. EPA. 1992b. The inorganic chemical characterization of water

treatment plant residuals. EPA/600/SR-92/117. Cincinnati, OH.

U.S. EPA. 1992c. Toxicity characteristic leaching procedure, method

1311. In: Test methods for evaluating solid waste, Vol. 1C. Laboratory manual--physical/chemical methods. SW-846, Revision O.

NTIS PB88239223. July.

U.S. EPA. 1993a. Quality assurance and quality control for waste

containment facilities. Technical guidance document. EPA/600/R93/182. Washington, DC.

U.S. EPA. 1993b. Sewage sludge sampling techniques (video).

Washington, DC: Office of Water. March.

U.S. EPA. 1993c. Large water system byproducts treatment and disposal cost document. Draft final report. EPA/811/D-93/002. Washington, DC.

U.S. EPA. 1994a. Design, operation, and closure of municipal solid

waste landfills. Seminar Publication. EPA/625/R-94/008. Washington, DC.

U.S. EPA. 1994b. Suggested guidelines for disposal of drinking water

treatment waste containing radioactivity (draft). Washington, DC:

Office of Groundwater and Drinking Water. June.

U.S. EPA. 1995a. Biosolids management handbook for small municipal wastewater treatment plants, U.S. Environmental Protection

Agency Regions 7, 8, and 10. Denver, CO.

U.S. EPA. 1995b. Process design manual: Land application of sewage sludge and domestic septage. EPA/625/R-95/001. Cincinnati,

OH.

USGS. 1982. Probabilistic estimates of maximum acceleration of velocity in rock in the contiguous United States. U.S. Geological

Survey Open-file Report 82-1033.

van Nieuwenhuyze, R.F., N.E. McTigue, and R.G. Lee. 1990. Beneficial applications and innovative sludge disposal methods. In:

Cornwell, D.A., and H.M.M. Koppers, eds. Slib, schlamm, sludge.

Denver, CO: American Water Works Association Research Foundation. pp. 267-304.

Vandermeyden, C., and E.K. Potter. 1991. Joint composting of water

and wastewater sludge. Presented at Joint AWWA/WPCF Residuals Management Conference, Research Triangle Park, NC.

Voorhees, J.R. 1974. Model studies on the effects of alum sludge on

the activated sludge waste treatment process. Master’s thesis.

Auburn University, Auburn University, AL.

Wang, M.C., J.Q. Hull, and M. Jao. 1992a. Stabilization of water

treatment plant sludge for possible use as embankment material.

Transportation Research Record 1345. Washington, DC: Transportation Research Board.

Wang, M.C., J.Q. Hull, M. Jao, B.A. Dempsey, and D.A. Cornwell.

1992b. Engineering behavior of water treatment sludge. J. Environ. Engin. ASCE 118(6).

Washington, H.G. 1984. Diversity, biotic and similarity indices: A review with special relevance to aquatic ecosystems. Water Resources Res. 18(6):653-694.

Watson, J.E., Jr., and D.J. Crawford-Brown. 1991. Use of activated

carbon to remove radon from drinking water. UNC-WRRTI-91-260.

Chapel Hill, NC: Water Resources Research Institute of the University of North Carolina.

W.E. Gates and Associates, Inc. 1981. A study of wastewater discharges from water treatment plants. Technical Report to the Ohio

River Valley Water Sanitation Commission.

Weaver, G. 1985. Sewering alum sludge: A look down the pipe. Public

Works.

Werle, C.P., J.T. Novak, W.R. Knocke, and J.H. Sherrard. 1984. Mixing intensity and polymer sludge conditioning. J. Environ. Eng.

Division of American Society of Civil Engineers. 110(5):191.

Westerman, A.G., and W.J. Birge. 1978. Accelerated rate of albinism

in channel catfish exposed to metals. Progressive Fish Culturist

40(4):143-146.

White, R.E. 1981. Retention and release of phosphate by soil and

soil constituents. Soil and Agriculture. Critical Report on Applied

Chemicals 2:71.

White, S.K., and E.A. Bondietti. 1983. Removing uranium by current

municipal water treatment processes. JAWWA 75(7):374-380.

Wilson, T.E., R.E. Bizzarri, T. Burke, P.E. Langdon, Jr., and C.M.

Courson. 1975. Upgrading primary treatment with chemicals and

water treatment sludge. J. Water Pollut. Control Fed.

47(12):2,820-2,833.

Wisconsin Department of Natural Resources. 1984. Radium in Wisconsin groundwater and removal methods for community water

systems.

Wisconsin Department of Natural Resources. 1985. Interim guidelines

for the disposal of liquid and solid wastes containing radium for

Wisconsin water treatment plants.

Witters, H., J.H.D. Vangenechten, S. Van Puymbroeck, and O.L.J.

Vanderborght. 1984. Interference of aluminum and pH on the

Na-influx in an aquatic insect Corixa punctata (Illig.) Bull. Environ.

Contam. Toxicol. 32:575-579.

Wood, C.M., and D.G. McDonald. 1987. The physiology of acid/aluminum stress in trout. In: Witters, H., and O. Vanderborght, eds.

Ecophysiology of acid stress in aquatic organisms. Soc. R. Zool.

Belg. 117:399-410.

Young, R.C., W.D. Burrows, and F.G. Ziegler. 1975. Biological survey

of Five Mile Creek, Jefferson County, Alabama. Associated Water

and Air Resources Engineers, Nashville, TN.

252
Zarini, S., D. Annoni, and P. Ravera. 1983. Effects produced by aluminum in freshwater communities studied by “enclosure” method.

Environmental Technol. Lett. 4:247-256.

253
Appendix A

Survey of State Regulatory Requirements: Summary of Results

The American Water Works Association’s (AWWA’s)

Water Treatment Residuals Management Committee

and the American Society of Civil Engineers (ASCE)

conducted a survey of state regulations governing the

disposal of water treatment plant residuals. The results

of this survey were presented at the 1991 Annual AWWA

Conference and Exposition Preconference Seminar,

“Current Perspectives on Water Treatment Plant Sludge

Disposal: Practices and Regulations” on June 23, 1991,

in Philadelphia, Pennsylvania (AWWA/ASCE, 1991).

Those results are summarized below.

ALABAMA

Department of Environmental Management

Municipal Waste Department

1751 Dickinson Drive

Montgomery, AL 36130

(205) 271-7825

The state regulates water treatment plant (WTP) wastes

under the Division of Municipal Wastes in the Department of Environmental Management. All discharges are

regulated and require a permit. The permits have limits

of waste constituents (liquid wastes) that must be monitored and cannot be exceeded. Most liquid wastes, such

as backwash decant, are sent to settling basins, then

discharged to surface water bodies after settling. Monofilling on site is the most common disposal practice for

sludges.

ALASKA

Department of Environmental Conservation

Treatment Section

P.O. Box O

Juneau, AK 99811-1800

(907) 465-2656

No regulations exist specifically covering WTP waste

disposal. A state statute requires a waste disposal permit to discharge solid or liquid waste to lands or waters

of the state. The statute also provides that a state-certified U.S. Environmental Protection Agency (EPA) National Pollutant Discharge Elimination System (NPDES)

permit also serves as a state waste disposal permit.

ARIZONA

Department of Environmental Quality (ADEQ)

P.O. Box 600

Phoenix, AZ 85001-0600

(602) 392-4002

ADEQ is responsible for the regulation of water plant

wastes within the state. ADEQ is in the process of

implementing the NPDES program and obtaining primacy from EPA. Currently, Arizona has specific regulations governing land application of wastewater biosolids;

these regulations are expected to apply to WTP residuals as well. Landfilling, monofilling, and co-disposal

have no additional state requirements beyond those

given in the existing federal regulations. The Department’s current policy allows the disposal of WTP residuals provided that 1) sampling and analysis of the

residuals reveal that they are nonhazardous; and 2) the

residuals are dewatered prior to disposal. No permits

are required by ADEQ for disposal of nonhazardous

WTP residuals.

ARKANSAS

Department of Pollution Control and Ecology

P.O. Box 8913

Little Rock, AR 72219-8913

(501) 570-2164

Arkansas has been granted the NPDES permit process

by EPA. They have a general permit that must be met

by all surface water treatment plants. Parameters regulated in the permit include aluminum, manganese, iron,

pH, and total suspended solids (TSS). If land application

is practiced, a permit is required. Co-disposal is commonly practiced in standard municipal landfills.

CALIFORNIA

Regional Water Quality Control Board

Los Angeles Region

101 Center Plaza Drive

Monterey Park, CA 91754

(213) 266-7512

California has regulations that classify wastes, including

water plant wastes, as hazardous, inert, and other cate254
gories. The classification system is used to identify the

appropriate management plan. The Regional Board can

regulate waste streams that are discharged to surface

water. A broad range of management practices is used,

including monofilling, co-disposal, direct discharge, and

land application. For WTP residuals to be landfilled, the

landfill must be equipped with a leachate collection and

removal system. The solids content must be at least 15

percent and a minimum solids-to-liquid ratio of 5 to 1 by

weight must be maintained. The state also maintains

criteria regarding the burial or the release of radioactive

material into air, water, or sanitary sewer systems.

COLORADO

Department of Public Health and Environment (CDPHE)

Water Quality Control Division

4300 Cherry Creek Drive South

Denver, CO 80222-1530

(303) 692-3500

CDPHE classifies WTP residuals and wastewater treatment plant (WWTP) biosolids according to heavy metal

and PCB concentrations. Classes 1 and 2 can be applied to agricultural land. Class 3 can be applied to

non-food chain land only. Class 4 is regulated by solid

waste rules. Monitoring requirements increase with the

higher classifications. Common disposal practices include monofilling, land application on site, lagooning,

and recycle/reuse. CDPHE permits the co-application of

WTP residuals and WWTP biosolids to land at a rate at

which the phosphorus fixing capability of the mixture

does not exceed the available phosphorus content of the

mixture.

CONNECTICUT

Not contacted.

DELAWARE

Delaware Pollution Control Branch

P.O. Box 1401

Dover, DE 19903

(302) 739-5731

Only two surface water treatment plants exist in the

state. Each discharge is reviewed and permitted separately according to NPDES guidelines.

FLORIDA

Department of Environmental Regulation (DER)

Drinking Water Section

2600 Blair Stone Road

Tallahassee, FL 32399-2400

(904) 487-1762

DER regulates water treatment plant wastes through the

Drinking Water Section. Florida is not a NPDES-delegated state and therefore is unable to issue NPDES

permits. The majority of wastes are lime softening or

reverse osmosis (RO) concentrates. RO concentrates

can be disposed of through ocean outfalls. Recycling

other wastewaters to the head of the plant is the most

common practice. Most residuals are recycled or landapplied, with some lime softening residuals used in road

bed mixtures. Monofills are also used. Radium content

in residuals is a major concern.

GEORGIA

Drinking Water Program

205 Butler Street Southeast

Floyd Tower East

Suite 1066

Atlanta, GA 30334

(404) 656-2750

Georgia regulates WTP wastes under the Drinking

Water Program with aid from the Wastewater Division of

NPDES. Georgia law requires all WTPs to have waste

disposal facilities and plans, and the Water Program

enforces all permits. No specific regulations address

WTP residuals. Most liquid wastes such as backwash

decant are sent to settling basins, then discharged to

surface water bodies after settling. Residuals are either

disposed of in municipal landfills (co-disposal) or by land

application.

HAWAII

Department of Health

Safe Drinking Water Branch

5 Waterfront Plaza, Suite 250-C

500 Ala Moana Boulevard

Honolulu, HI 96813

(808) 543-8258

Hawaii is aware of the concern about WTP residuals,

but because of staffing and budget constraints, it is

unable to actively pursue a waste treatment program.

Land application is the primary method of disposal. The

state is awaiting EPA guidance.

IDAHO

Division of Environmental Quality (DEQ)

Water Quality Bureau

1410 North Hilton

Boise, ID 83706

(208) 334-5865

DEQ does not administer the NPDES program, and

there are no plans for developing regulations for WTP

residuals at this time. The most common disposal practice is probably landfilling.

255
ILLINOIS

Environmental Protection Agency

2200 Churchill Road

Springfield, IL 62794-9276

(217) 785-8653

Illinois EPA regulates WTPs that discharge any treatment waste stream to waters of the state. The pollutants

regulated depend on the plant, but suspended solids,

pH, and iron are the primary concerns. The most common method of disposal is discharge to sanitary sewers.

Lagooning is also used, with effluent discharged to

streams (NPDES permit required) and solids disposed

of in a landfill. Another concern is the discharge of salt

backwash from zeolite softeners to sanitary sewers.

INDIANA

Department of Environmental Management

105 South Meridian Street

P.O. Box 6015

Indianapolis, IN 46206

(317) 233-4222

WTP residuals have not been a major problem for the

state. Filter backwash is usually discharged to a sanitary

sewer or allowed to percolate through the ground.

Sludge solids from surface water plants are usually land

applied. The state is responsible for administering the

NPDES program.

IOWA

Department of Natural Resources (DNR)

Environmental Protection Division

Wastewater Permit Section

900 East Grand

Des Moines, IA 50319

(515) 281-8998

DNR is seeking authority from the state legislature to

develop a permit by rule process in which permit requirements would be identified by a DNR rule. Discharges to

surface waters require an NPDES permit. Land application of WTP residuals may be regulated similarly for

industrial and sewage sludges with respect to metals.

WTP residuals have not been a primary concern because of limited resources.

KANSAS

Department of Health and Environment

Bureau of Water

Forbes Field

Topeka, KS 66620

(913) 296-5503

The state regulates all types of water wastes but has not

actively pursued permitting facilities. At this time, the

primary objective is to keep softening sludge solids out

of the streams and rivers. Rivers have been identified

as nonattainment areas for suspended solids by nonpoint legislation. In the future, this may affect management of presedimentation solids. The state administers

the NPDES program for discharges; lagoons are permitted through a state water pollution control permit.

KENTUCKY

Division of Waste Management

Solid Waste Branch

18 Reilly Road

Frankfort, KY 40601

(502) 564-3410

The state regulates WTP residuals under the Division of

Solid Waste. While specific regulations are being written, disposal is presently handled on a case-by-case

basis. Some liquid wastes, such as backwash decant,

are sent to settling basins, then discharged to surface

water bodies after settling, while others are recycled to

the head of the plant. Roughly half of the plants land-apply sludges, and half send the solids to a landfill (co-disposal). A few monofills exist.

LOUISIANA

Department of Environmental Quality

Water Pollution

P.O. Box 82215

Baton Rouge, LA 70884-2215

(504) 765-0635

WTP residuals are regulated by the state’s water pollution control programs. An NPDES permit, administered

jointly by the state and EPA, is required for surface water

discharges. Only liquid wastes may be discharged. Solid

wastes are commonly disposed of in permitted landfills.

MAINE

Department of Environmental Protection (DEP)

(207) 582-8740

The DEP requires that all WTP residuals be dewatered

to a minimum of 20 percent total solids and disposed of

in a secure landfill. The landfills must be equipped with

a system for leachate collection.

MARYLAND

Not contacted.

MASSACHUSETTS

(617) 292-5529

The state is in the process of updating its existing regulations. Sludge disposal is considered on a case-bycase basis. The state is concerned about some gaps in

the existing regulations.

256
MICHIGAN

Department of Public Health (DPH)

3423 North Logan

Lansing, MI 48909

(517) 335-9216

The DPH requires water utilities to develop a management plan for WTP residuals. If the plan is acceptable,

no further regulation is necessary. If backwash water is

discharged to a surface water, an NPDES permit is

required from the Department of Natural Resources.

The most common methods of disposal include co-disposal, direct discharge, and land application.

MINNESOTA

Not contacted.

MISSISSIPPI

Office of Pollution Control

P.O. Box 10385

Jackson, MS 39289-0385

(601) 961-5171

Mississippi regulates WTP residuals under the Office of

Pollution Control. New regulations take their cue from

EPA and use federal effluent guidelines for wastes, particularly with regard to toxicity. Most liquid wastes, such

as backwash decant, are sent to settling basins, then

discharged to surface water bodies after settling. A few

plants discharge to a publicly owned treatment works

(POTW) or recycle washwater to the head of the plant.

The majority of sludges go to municipal landfills, with

some land application of waste solids. No monofills exist

in the state.

MISSOURI

Department of Natural Resources

P.O. Box 176

Jefferson City, MO 65102

(314) 751-6624

WTP residuals are regulated as solid waste unless the

material is applied to agricultural land or is discharged

to a surface water. A solid waste permit may be required

for landfill disposal, which could include requirements

for TCLP testing, liners, and leachate collection systems. An NPDES permit is required for direct discharge.

Solids may be discharged only into large rivers; treated

filter backwash may be discharged to small rivers. No

restrictions are placed on land application of lime softening sludges. Guidelines for total aluminum loadings

to the soil are established for the application of alum

sludges. With some limitations, brine wastes may be

discharged to a sanitary sewer, abandoned quarries,

gravel pits, or evaporation lagoons.

MONTANA

Department of Health and Environmental Services

(MDHES)

Water Quality Building

Cogswell Building

Helena, MT 59620

(406) 444-2406

MDHES may regulate all types of WTP residuals on a

case-by-case basis. In the future, these residuals may

be addressed along with sewage sludges. NPDES regulations apply to any discharges. The majority of the

sludges are from lime softening facilities and are landapplied. The primary concern of this practice is checking

soil pH to ensure suitability for application.

NEBRASKA

Department of Environmental Control (NDEC)

P.O. Box 98922

Lincoln, NE 68509-8922

(402) 471-2186

NDEC regulates disposal of WTP residuals on a caseby-case basis. In the future, residuals may be addressed

along with sewage sludges. NPDES regulations apply

to any discharges. The majority of the sludges are from

lime softening facilities and are land-applied.

NEVADA

Division of Environmental Protection

123 West Nye Lane

Carson City, NV 89710

(702) 687-5872

In general, WTP residuals are not a major concern

because they are nonhazardous. No permits are required for landfilling solid waste; however, discharge of

treated washwater to rivers requires a permit. The state

plans to update existing regulations in the near future

and to develop new regulations within 1 to 5 years.

NEW HAMPSHIRE

(603) 271-3503

The state encourages landfilling and land application of

WTP residuals. The state plans to update existing regulations within 1 year. Concerns include heavy metals

and nutrients.

NEW JERSEY

(609) 292-5550

The state regulates WTP residuals treatment and disposal processes on a case-by-case basis and is responsible for administering the NPDES program. Concerns

include heavy metals and organics.

257
NEW MEXICO

Environmental Department

P.O. Box 26110

Santa Fe, NM 87502

An NPDES permit is required for any discharges from

water treatment plants. The EPA has primacy for permitting discharges. The state receives a 106 grant from the

Clean Water Act to assist EPA with compliance monitoring. The most common methods of disposal include

co-disposal and direct discharge.

NEW YORK

Department of Environmental Conservation

(518) 457-1067

Permits are issued based on the type of disposal practice proposed and the sludge quality characteristics. A

wide range of disposal practices, including monofilling,

co-disposal, direct discharge, land application, and recycle/reuse, are commonly used. Concerns include

heavy metals, organics, and nutrients.

NORTH CAROLINA

Division of Environmental Management

P.O. Box 29535

Raleigh, NC 27626

(919) 733-5083

North Carolina regulates WTP residuals through the

Division of Environmental Management. Wastes are

classified by the treatment process, and a certified operator is required to run the waste treatment aspects of

the plant. Most liquid wastes are discharged to a POTW

via a sanitary sewer. Some direct discharge, recycling

to the head of the plant, and subsurface disposal of

liquid wastes are practiced. Solids disposal is generally

through a municipal landfill, although there are a few

monofills and some land application of sludges.

NORTH DAKOTA

Department of Health

Solid Waste Division

P.O. Box 5520

Bismarck, ND 58502-5520

(701) 221-5166

The state has no specific rules for WTP residuals. Technically, all types of water treatment plant residuals are

regulated; for the most part, only lime softening wastes

are regulated. Solids are disposed of primarily by landfilling, which is regulated on a case-by-case basis. Landfills must have a plan established for managing the

waste.

OHIO

Environmental Protection Agency

Division of Public Drinking Water

18 West Water Mark Drive

Columbus, OH 43266

(614) 644-2752

The state is responsible for administering the NPDES

program. Discharge of backwash water is regulated by

water quality standards and limitations have been established for TSS, pH, iron, manganese, and chlorides. The

state prefers that utilities discharge to a sanitary sewer.

No unregulated discharges exist, and wastes are not

recycled through the plant. The primary disposal method

for solids is agricultural land application of lime softening

sludge. Other acceptable methods of disposal include

landfilling and mixing with compost.

OKLAHOMA

State Department of Health

1000 Northeast 10th Street

Oklahoma City, OK 73152

(405) 271-7370

Oklahoma jointly administers the WTP residuals program with EPA. Permits for direct discharges require

analysis for pollutants listed in 40 CFR Part 122, Appendix D, Tables II and III. The most common disposal

practices include co-disposal, direct discharge, and land

application. WTP residuals are being more closely controlled.

OREGON

Department of Environmental Quality

811 S.W. Sixth Avenue

Portland, OR 97204

(503) 229-5782

Disposal of WTP residuals in Oregon is regulated under

solid waste management rules. A solid waste permit is

required to construct and operate onsite waste handling

facilities. Offsite agricultural land application practices

are covered in the permit. For discharge to surface

water, an NPDES permit is required. The most common

methods of disposal include monofilling and co-disposal

with solid waste.

PENNSYLVANIA

Department of Environmental Resources

Room 518, Executive House, South 2nd Street

Harrisburg, PA 17105

(717) 787-9035

The state has primacy for regulating WTP residuals. It

regulates backwash solids, alum, ferric chloride, polymer, and activated carbon discharge. The primary concern has been solids, alum, and ferric wastes. (No

258
softening plants exist in the state.) The NPDES classification system is used with organics, inorganics, TSS,

and pH monitored. Common sludge disposal includes

landfill and land application. Backwash water can be

recycled to the head of the plant or lagoon-treated. No

new regulations are planned.

RHODE ISLAND

Not contacted.

SOUTH CAROLINA

Department of Health and Environmental Control

2600 Bull Street

Columbia, SC 29201

(803) 734-5241

The state regulates WTP residuals under the Industrial

Wastewater Division of the Department of Health and

Environmental Control. Reverse osmosis concentrate

has recently become a concern. Most liquid wastes,

such as backwash decant, are sent to settling basins,

then discharged to surface water bodies after settling.

Few facilities discharge to POTWs. Future goals are to

encourage recycling of all washwaters to the head of the

plant. Most sludges are disposed of at municipal landfills

or transferred to a WWTP for mixing and disposal with

sewage sludge. Land application of sludges is not

practiced.

SOUTH DAKOTA

Department of Environment and Natural Resources

Office of Drinking Water

Joe Foss Building, Room 412

Pierre, SD 57501

(605) 773-3754

Landfilling and land application are the primary methods

of WTP residuals disposal. Therefore, the solid waste

program covers most of the monitoring requirements.

Landfills must be permitted. Radium removal is a concern, with allowable limits set at the NRC’s limits for

low-level radioactive wastes.

TENNESSEE

Water Quality Control Department

Industrial Facilities Section

150 Ninth Avenue North, 4th Floor

Nashville, TN 37247-3001

(615) 741-7883

Tennessee regulates WTP residuals under the Industrial

Facilities Section of the Water Quality Control Department. New regulations will be developed within the year

for filter backwash wastes. The most common disposal

practice for liquid wastes is to send them to a POTW,

while discharge to surface water bodies and recycling to

the head of the plant are lesser used options. Sludges

are most often disposed of in municipal landfills, yet a

few plants land-apply or monofill their solid wastes.

TEXAS

Department of Environmental Health

1100 West 49th Street

Austin, TX 78756

(512) 458-7542

Permitting is performed jointly with EPA, but the Texas

Water Commission controls the permitting. Treated effluent from filter backwash and clarifier blowdown is

typically recycled to the plant. Waste solids are usually

not discharged to the watercourse. Sludge disposal is

regulated in two ways: 1) onsite disposal must be registered with the Municipal Waste Department, and 2) offsite disposal requires a solid waste permit from the

Department.

UTAH

Bureau of Water Pollution Control

P.O. Box 16690

Salt Lake City, UT 84116

(801) 538-6146

The most common disposal practices for WTP residuals

are landfilling and land application. The state has no

specific regulations for WTP residuals. Best management practices are required for disposal. These practices include containment (no runoff), dewatering, and

disposal in a way that does not create a nuisance. No

discharge is allowed to natural drainage.

VERMONT

(802) 244-7831

The most common options for disposal of WTP residuals

include landfilling and land application. All types of water

plant wastes are regulated. The constituents of concern

include heavy metals, organics, and nutrients.

VIRGINIA

Department of Health

State Water Control Board

(804) 786-1766

The Department of Health regulates backwash water

reuse. Direct return of the backwash to the treatment

train is discouraged. The State Water Control Board

monitors and permits plants on a case-by-case basis.

Constituents of concern for discharges include aluminum, TSS, pH, and chlorine residual.

259
WASHINGTON

Department of Ecology

Mail Stop PV-11

Olympia, WA 98504-8711

(206) 438-7037

The state has established guidelines for the disposal of

WTP residuals. The guidelines establish NPDES permit

requirements, including effluent limitations and monitoring requirements, and guidelines for permitting mechanical dewatering and sewage disposal facilities. A state

statute allows plants on specific rivers to discharge sediment to the originating river. EPA may prohibit this practice by not concurring with the draft NPDES permit.

WEST VIRGINIA

Department of Natural Resources (DNR)

Water Resources Division

1201 Greenbriar Street

Charleston, WV 25375

(304) 348-2107

DNR is responsible for administering the NPDES program. Treated filter backwash liquid requires a permit for

discharge to rivers. Discharge criteria are established on

a case-by-case basis but typically include criteria for

aluminum and other metals. Solids are typically land

applied, but some sewer discharge is practiced. WTP

residuals have not been a primary concern for the state.

WISCONSIN

Department of Natural Resources

(608) 266-2290

Wisconsin is in the process of upgrading all regulations.

The state requires a complete analysis of the waste

stream, and an NPDES permit is required for discharge

to surface waters. For backwash water, the state prefers

pond percolation into the ground. This practice requires

a ground-water discharge permit. The primary method

of waste disposal is discharge to a sanitary sewer system.

WYOMING

Department of Environmental Quality

Water Quality Division

Herscher Building, 4 West

Cheyenne, WY 82002

(307) 777-7781

WTP residuals have not been a major concern. No

discharges to natural drainage are allowed, and all filter

backwash effluent must be recycled to the plant. Landfilling is the most common disposal practice and is regulated by the solid waste criteria, which primarily require

dewatering.

260
Appendix B

1992 Survey of Water Treatment Plants Discharging to 

Wastewater Treatment Plants

All of the water treatment plants that appear in the following tables (AWWA/AWWARF, 1992) treat surface water.

Size of Plant Chemicals Added at WTP

Design

Capacity

(MGD)

Average

Daily

Demand 

(MGD)

Maximum 

Daily

Demand

(MGD)

Peak

Hourly

Demand

(MGD)

Major Processes

at Water

Treatment Plant

Coagulant

(lb/day)

Coagulant 

Aid

(lb/day)

Filter

Aid

(lb/day)

PAC

(lb/day)

Oxidant/

Disinfectant

(lb/day)

Other

(lb/day)

F.E. Weymouth

Filtration Plant,

LaVerne, CA,

Metropolitan Water

District

520 330 — — Coagulation,

filtration

Alum 7,930 Cationic

polymer

2,420

— — Chlorine

8,120

Ammonia 

750

Robert A. Skinner

Filtration Plant,

Temecula, CA,

Metropolitan Water

District

520 300 402 402 Coagulation,

filtration

Alum 15,290 Cationic

polydadmac

3,290

— — Chlorine

9,935

Ammonia 

1,135

Joseph Jensen

Filtration Plant,

Granada Hills, CA,

Metropolitan Water

District

550 333 — — Coagulation,

filtration

Alum 5,510 Cationic

polydadmac

2,770

Polyacrylamide 28— Chlorine

8,720

Ammonia 

800

Dunkirk WTP,

Dunkirk, NY

8 4 — — Coagulation, lime

softening, filtration,

carbon adsorption

(summer)

Polyaluminum

chloride 100

— — 500–1,000 

(July–Aug)

Chlorine 65 Lime 30

Chesterfield County

WTP

12 10.5 — — Fe/Mn removal,

coagulation, lime

softening, filtration

City of Myrtle

Beach WTP, NC

29.5 14 20 1 Coagulation,

filtration, ozone

Alum

13,000–2,000

— 20–3 — Ozone

2,800–400

Chlorine

2,000–500

261
Residuals Streams

at WTP

Monitored Characteristics

of Residuals Streams Residuals Handling Facilities

Filter

Backwash

Clarification

Basin

Sludge Other TSS

COD/ 

BOD pH Metals Other

Equalization

Basins

Chemical

Recovery

Process

Streams

Recycled

DewateringThickening

Final %

Solids at

Treatment 

Plant

F.E. Weymouth

Filtration Plant,

LaVerne, CA,

Metropolitan

Water District

31.7

MGD

1.4 

MGD

— Yes COD Yes — Sulfides — — Yes Cationic

polyacrylamide

— Yes <3%

Robert A.

Skinner Filtration

Plant, Temecula,

CA, Metropolitan

Water District

7.5

MGD

0.036

MGD

— Filter

backwash 

3%

— — Yes Yes Yes, belt

press

Yes <1%

Joseph Jensen

Filtration Plant,

Granada Hills,

CA, Metropolitan

Water District

21

MGD

(max)

0.8

MGD

(max)

— Yes — — Future

construction to

pump to

head of

plant

Cationic

polyacrylamide

— Yes 2-3%

Dunkirk WTP,

Dunkirk, NY

Yes Yes — — Yes,

holding

tank

— — — — — 0.5–1.0%

Chesterfield

County WTP

325,000 

GPD

— — 250 Yes — — — — — 16%

City of Myrtle

Beach WTP, NC

0.3

MGD

0.043

MGD

— — — — Backwash

decant

— Constructing mechanical

dewatering 

processes

— 1%

Conveyance to WWTP

Where Residuals

Introduced

Problems

Encountered

Formal Agreement

With WWTP

Alternate Disposal/Beneficial

Reuse Options

F.E. Weymouth Filtration Plant,

LaVerne, CA, Metropolitan

Water District

Discharge to sewer Front end — Surcharge if exceed

levels

Yes

Robert A. Skinner Filtration

Plant, Temecula, CA,

Metropolitan Water District

Discharge to sewer Front end — — This WTP has onsite

reclamation plant, full liquid

recovery system, and onsite

disposal approved site

Joseph Jensen Filtration Plant,

Granada Hills, CA, Metropolitan

Water District

Discharge to sewer Front end — Heavy metals not

accepted; surcharge 

if exceed levels

Yes

Dunkirk WTP, Dunkirk, NY Pumped through force main Front end — — —

Chesterfield County WTP Pumped through force main Front end — — Land application

City of Myrtle Beach WTP, NC Pumped through force main Front end decant 

from alum lagoons

— — Composting

262
Size of Plant Chemicals Added at WTP

Design

Capacity

(MGD)

Average

Daily

Demand 

(MGD)

Maximum

Daily

Demand

(MGD)

Peak

Hourly

Demand

(MGD)

Major Processes

at Water

Treatment Plant

Coagulant

(lb/day)

Coagulant 

Aid

(lb/day)

Filter

Aid

(lb/day)

PAC

(lb/day)

Oxidant/

Disinfectant

(lb/day)

Other

(lb/day)

City of Greensboro

WTP, NC

44 31 44 54 Coagulation,

filtration

Alum 5,900 — — — Chlorine

500

—

Texarkana WTP, TX 18 11 17 Sedimentation,

filtration

— — — — Chlorine

dioxide

chloramines

Potassium

permanganate

Knoxville WTP, TN 60 30–33 40–42 34 Coagulation,

filtration

Alum 600 — 20 — Chlorine

1,200

—

Nashville WTP, TN 180 120 140 — Coagulation,

softening, filtration 

Alum — — — — Lime

Queens Lane WTP,

Philadelphia, PA,

City of Philadelphia

100 100 120 130 Coagulation,

filtration

Ferric

chloride;

alum as

needed

— — PAC as

required

— Lime

Belmont WTP,

Philadelphia, PA,

City of Philadelphia

80 60 75 90 Coagulation,

filtration

Ferric

chloride;

alum backup

— — PAC as

required

— Lime

Franklin WTP, NC,

Charlotte-Mecklenburg

Utility District

96 47.6 71.4 — Coagulation,

filtration

Alum 

9.0 mg/L

— — Carbon

0.9 mg/L

— —

Vest WTP, NC,

Charlotte-Mecklenburg

Utility District

24 18.9 24 — Coagulation,

filtration

Alum 

9.2 mg/L

— — Carbon

<1.0

mg/L

— —

City of Boulder

WTP, CO

12 5.3 12 12 Coagulation,

filtration

Alum — — Yes Prechlorination—

T.W. Moses WTP,

Indianapolis, IN,

Indianapolis Water

Company

16 10 20 1 Coagulation,

filtration

Alum Cationic

polymer

— — Chlorine

263
Residuals Streams

at WTP

Monitored Characteristics of

Residuals Streams Residuals Handling Facilities

Filter

Backwash

Clarification

Basin

Sludge Other TSS

COD/ 

BOD pH Metals Other

Equalization

Basins

Chemical

Recovery

Process

Streams

Recycled Chemical

Dewatering

Thickening

Final %

Solids

at Treatment

Plant

City of Greensboro

WTP, NC

0.5

MGD

0.1

MGD

— Not available Yes Yes —

Texarkana WTP, TX Yes — — 300–700

mg/L

— — — — Filter backwash and

sedimentation basins

Knoxville WTP, TN — Yes — 30,000

mg/L

— — — — — — — — — Gravity

thickener

2%

thickened 

solids

Nashville WTP, TN Not available Yes — — — — — —

Queens Lane WTP,

Philadelphia, PA City

of Philadelphia

Yes Yes — Not available — — — — — — —

Belmont WTP, City of

Philadelphia, PA

Yes Yes — Not available — — — — — — —

Franklin WTP, NC

Charlotte-Mecklenburg

Utility District

Yes Yes — — — — Yes Yes Yes — — — — — —

Vest WTP, NC

Charlotte-Mecklenburg

Utility District

Yes Yes — — — — Yes Yes — — — — — — —

City of Boulder WTP,

CO

Yes Yes — — — — — — Yes — Backwash

water

— —— —

T.W. Moses WTP,

Indianapolis, IN

Indianapolis Water

Company

Yes Yes — — — — — — — — Yes — — — —

264
Conveyance to

WWTP

Where Residuals

Introduced Problems Encountered

Formal Agreement

With WWTP

Alternate

Disposal/Beneficial

Reuse Options

City of Greensboro WTP, NC Discharge to sewer Front end — WTP and WWTP

operated by same

agency

—

Texarkana WTP, TX Discharge to sewer Front end — WTP and WWTP

operated by same

agency

—

Knoxville WTP, TN Pumped through

force main

Front end Problems in flocculation

basin; a lot of sedimentation;

yearly maintenance

WTP and WWTP

operated by same

agency

—

Nashville WTP, TN Pumped through

force main

Front end; influent

chamber

Minor problems on small line — Composting

Queens Lane WTP,

Philadelphia, PA, City of

Philadelphia

Discharge to sewer Front end WWTP can handle routine

discharges but seasonal

basin cleaning causes

problems

Discharge restricted

by permit

Evaluating onsite

dewatering with disposal

to landfill/reclamation on

site

Belmont WTP, Philadelphia, PA,

City of Philadelphia

Discharge to sewer Front end WWTP can handle but

seasonal basin cleaning at

WTP must be cleared first

Discharge restricted

by permit

Evaluating onsite

dewatering with disposal

to landfill/reclamation on

site

Franklin WTP, NC,

Charlotte-Mecklenburg Utility

District

Discharge to sewer Front end Pretreatment and toxicity

concerns at WWTP; this

disposal method will not be

feasible in future due to

requirements of discharge

permits for WWTP

WTP and WWTP

operated by same

utility

Co-disposal with biosolids;

agronomic land

application; dedicated land

application; turf farming

Vest WTP, NC,

Charlotte-Mecklenburg Utility

District

Discharge to sewer Front end Pretreatment and toxicity

concerns at WWTP; this

disposal method will not be

feasible in future due to

requirements of discharge

permits for WWTP

WTP and WWTP

operated by same

utility

Co-disposal with biosolids;

agronomic land

application; dedicated land

application; turf farming

City of Boulder WTP, CO Discharge to sewer Front end — — Co-disposal with

wastewater sludge

T.W. Moses WTP, Indianapolis,

IN, Indianapolis Water Company

Gravity main Front end Main problem encountered is

too thick a sludge for gravity

flow

Allowed a no-charge

level for

ammonia-nitrogen,

TSS, and BOD. The

WTP always exceeds

the levels for TSS and

BOD and never

exceeds nitrogen

allowance. There are

no restrictions

—

265
Size of Plant Chemicals Added at WTP

Design

Capacity

(MGD)

Average

Daily

Demand

(MGD)

Maximum 

Daily

Demand

(MGD)

Peak

Hourly

Demand

(MGD)

Major Processes at

Water Treatment

Plant

Coagulant

(lb/day)

Coagulant 

Aid

(lb/day)

Filter

Aid

(lb/day)

PAC

(lb/day)

Oxidant/

Disinfectant 

(lb/day)

Other 

(lb/day)

Nottingham WTP,

Cleveland, OH 

City of Cleveland

100 87.1 131 152 Disinfection,

adsorption, oxidation,

coagulation,

flocculation,

sedimentation,

filtration, fluoridation

Alum — — PAC Chlorine Sodium

silicofluoride,

potassium

permanganate

Morgan WTP,

Cleveland, OH 

City of Cleveland

150 80 140 150 Disinfection,

adsorption, oxidation,

coagulation,

flocculation,

sedimentation,

filtration, fluoridation

Alum — — PAC Chlorine Hydrofluosilicic

acid

Baldwin WTP,

Cleveland, OH 

City of Cleveland

165 85 155 162 Disinfection,

adsorption, oxidation,

coagulation,

flocculation,

sedimentation,

filtration, fluoridation

Alum — — PAC Chlorine Sodium

hypochlorite,

sodium

silicofluoride,

potassium

permanganate

Erie City Water

Authority, Erie, PA

60–80 35–40 60 Coagulation, filtration Polyalum

chloride

— — PAC Potassium

permanganate

Conestoga WTP,

Lancaster, PA 

City of Lancaster

16 7 10.2 0.425

MGH

Coagulation, filtration Alum — — PAC Chlorine Lime

(hydrated),

liquid

hydrofluosilicic

acid,

potassium

permanganate

Residuals Streams

at WTP

Monitored Characteristics of

Residuals Streams Residuals Handling Facilities

Filter

Backwash

Clarification

Basin

Sludge Other TSS

COD/ 

BOD pH Metals Other

Equalization

Basins

Chemical

Recovery

Process

Streams

Recycled Chemical

Dewatering

Thickening

Final %

Solids at

Treatment 

Plant

Nottingham WTP,

Cleveland, OH,

City of Cleveland

Yes Yes — 2,500–

4,000 

mg/L

BOD

210

mg/L

7.0–7.4 — — Attenuation

tank

— Yes — — — —

Morgan WTP,

Cleveland, OH,

City of Cleveland

Yes Yes — 2,500–

4,000 

mg/L

BOD

200

mg/L

7.0–7.4 — — — — Yes — — — —

Baldwin WTP,

Cleveland, OH,

City of Cleveland

Yes Yes — 2,500–

4,000 

mg/L

BOD

210

mg/L

7.0–7.4 — — — — Yes — — — —

Erie City Water

Authority, Erie, PA

Yes Yes — 30–60 2–14 7.6–8.0 — Alum,

75

mg/L

Yes — Yes — — Yes —

Conestoga WTP,

Lancaster, PA,

City of Lancaster

Yes Yes — — — — — — — — — — — Sludge,

partially

thickened

—

266
Conveyance to WWTP

Where Residuals

Introduced Problems Encountered

Formal Agreement

With WWTP

Alternate Disposal/Beneficial

Reuse Options

Nottingham WTP,

Cleveland, OH,

City of Cleveland

Solids diluted if necessary

and pumped through force

main

Front end — Yes —

Morgan WTP, Cleveland,

OH, City of Cleveland

Force main Front end Maintenance of optimal

discharge parameters for

cost-effectiveness requires

ongoing effort

Yes —

Baldwin WTP, Cleveland,

OH, City of Cleveland

Diluted then discharged to

sewer

Front end — Yes —

Erie City Water Authority,

Erie, PA

Discharge to sewer Front end — — —

Conestoga WTP,

Lancaster, PA,

City of Lancaster

Discharge to sewer Front end — — Considering beneficial use via

private hauling to Susquehanna

WTP, Lancaster, PA

Size of Plant Chemicals Added at WTP

Design

Capacity

(MGD)

Average

Daily

Demand 

(MGD)

Maximum 

Daily

Demand

(MGD)

Peak

Hourly

Demand

(MGD)

Major Processes

at Water

Treatment Plant

Coagulant

(lb/day)

Coagulant 

Aid

(lb/day)

Filter

Aid

(lb/day)

PAC

(lb/day)

Oxidant/

Disinfectant

(lb/day)

Other

(lb/day)

City of Milwaukee

Waterworks,

Milwaukee, WI

250 Winter–

60 

Summer–

120

171 240 Coagulation,

filtration

Poly-aluminum

hydroxychloride

— — PAC

when

needed

Chlorine Ammonia,

fluosilicic

acid

Chattanooga WTP,

TN

72 38 — 48 Coagulation,

filtration, GAC

Polymers — — — Chlorine Fluorine

Residuals Streams

at WTP

Monitored Characteristics of

Residuals Streams Residuals Handling Facilities

Filter

Backwash

Clarification

Basin

Sludge Other TSS

COD/

BOD pH Metals Other

Equalization

Basins

Chemical

Recovery

Process

Streams

Recycled Chemical

Dewatering

Thickening

Final %

Solids at

Treatment 

Plant

City of

Milwaukee

Waterworks,

Milwaukee, WI

Yes Yes Spent 

PAC

4,257

mg/L

310

mg/L

— — — — — Yes — — — —

Chattanooga

WTP, TN

Yes Yes — — —

Conveyance to

WWTP

Where Residuals

Introduced

Problems

Encountered

Formal Agreement 

With WWTP

Alternate Disposal/Beneficial

Reuse Options

City of

Milwaukee

Waterworks,

Milwaukee, WI

Discharge to sewer Front end Sometimes exceed

solids limit;

sometimes need to

flush sewer

Industrial surcharge;

discharge restrictions;

cannot exceed 20,000 lbs

solids per day:

Copper – 6.0 mg/L, Lead –

2.0 mg/L, Nickel – 4.0 mg/L,

Zinc – 8.0 mg/L

Chattanooga

WTP, TN

267
Appendix C

Charges From Publicly Owned Treatment Works

The publicly owned treatment works (POTW) charges

presented in Chapter 11 were based on a limited sampling of POTW rates conducted for this analysis and on

a survey published by the League of Minnesota Cities.

Table C-1 presents the results of the telephone sample,

and Table C-2 presents selected results from the

League of Minnesota Cities Survey. Ten of the cities

contacted during the telephone sample do not charge to

accept wastes from water treatment plants; three were

not certain. In addition, some of the POTWs contacted

during the telephone sample and by the League of

Minnesota Cities charge a flat rate per month with no

additional charges for higher flow rates.

Table C-1. Sewer Rates—Large Cities (DPRA, Inc., 1992)

City State Population Sewage Fee Charged BOD TSS Misc.

Little Rock Arkansas 175,818 $1.42/1,000 gallons +

monthly service charge

>300 ppm:

$0.092/pound

>300 ppm:

$0.0883/pound

Oil and grease

>100 ppm:

$0.047/pound

Tucson Arizona 405,390 Varies per industry a CODa

Chula Vista California 138,000 $16.70/month 0–200 ppm:

$1.97/1,000 gallons

200–499 ppm:

$2.42/1,000 gallons

>500 ppm:

$3.26/1,000 gallons

Denver Colorado 467,610 $1.65/1,000 gallons b b Pretreatment:

$450/million

gallons for

metal treating

facilities

$250/million

gallons for

other significant

industries 

Derby Connecticut 12,346 $0.82/gallon

(commercial)

$0.74/gallon (industrial)

No surcharge No surcharge

Boca Raton Florida 49,505 $1.70/1,000 gallons No surcharge No surcharge $0.25/gallon

surcharge

outside city

limits

Lewiston Idaho 27,986 $1.47/1,000 gallons $0.12/pound $0.11/pound Industry

surcharge:

$0.09/1,000

gallons

Anderson Indiana 64,695 $2.57/1,000 gallons >300 ppm:

$0.14/pound

>300 ppm:

$0.12/pound

Billing charge:

$0.77/month

Ft. Scott Kansas 8,893 2,244 gallons:

$6.16/month

2,244+ gallons:

$1.18/1,000 gallons

No surcharge No surcharge

Bloomington Minnesota 81,831 $1.53/1,000 gallons No surcharge No surcharge

268
City State Population Sewage Fee Charged BOD TSS Misc.

Duluth Minnesota 85,000 $4.25 + $2.32/1,000

gallons

cc c

Minneapolis Minnesota 370,951 $2.21/1,000 gallons >250 ppm:

$0.106/pound

MPCA charge

for

non-residential:

$25/month

COD

>250 ppm:

$0.053/pound

Bozeman Montana 24,500 $3.95 + $0.87/1,000

gallons

No surcharge No surcharge

Fargo North Dakota 76,000 $0.70/1,000 gallons

(Minimum: $7.70/month)

No surcharge No surcharge

Salem Oregon 100,000 $0.41/1,000 gallons $0.15/pound $0.117/pound $1.55 billing

charge;

$108/diameter

inch mile

Richmond Virginia 219,214 $1.79/1,000 gallons >250 ppm:

$12.82/

hundredweight

>275 ppm: 

$14.84/

hundred-weight

Seattle Washington 493,846 $3.42/1,000 gallons >300 ppm:

$0.093/pound

>400 ppm:

$0.14/pound

Metals:

$0.000806/

gallon

Oils, grease,

fats:

$0.000124/gallon

a If average TSS and COD are exceeded, a high strength factor ranging from $1.00 to $3.78 is assessed.

b (Vc) 8.34 [(0.142)(BOD–220)] + [(0.10)(TSS–250)]. c

 Surcharges unknown.

Key

BOD = Biological oxygen demand.

TSS = Total suspended solids.

COD = Chemical oxygen demand.

Table C-1. Sewer Rates—Large Cities (Continued)

Table C-2. Sewer Rates—Minnesota Cities (League, 1991)

City State Population Sewage Fee Charged Misc.

Benson Minnesota 3,656 $2.47/1,000 gallons

Wadena Minnesota 4,699 $1.55/1,000 gallons Minimum: $6.23

Ely Minnesota 4,820 $2.62/1,000 gallons Minimum: $7.84

Connection charge:

$75 digging fee 

Chisholm Minnesota 5,930 $0.90/1,000 gallons

Arden Hills Minnesota 8,012 $1.64/1,000 gallons

Andover Minnesota 9,387 Varies for “area”; ranges between $4.50

and $8.50/month

Bemidji Minnesota 10,949 Flat rate ranges between $4.45 and

$44.00

Owatonna Minnesota 18,632 $2.50/month + $1.78/1,000 gallons Minimum: $2.50

Albert Lea Minnesota 19,200 $3.65/month $14 connection charge

if turned off

Mankato Minnesota 28,651 $0.86/unit

Burnsville Minnesota 35,674 Commercial: 90% of winter water usage

or metered water

Connection charge:

$168/unit

Duluth Minnesota 92,811 $3.75/month + $1.72 to $2.99/1,000

gallons

269
Appendix D

Chemical Monthly Average Doses, Pine Valley Water Treatment Plant, 

Colorado Springs, CO, 1987–1992 (Pine Valley, 1994)

Table D-1. Raw Water, Pine Valley, 1987

Water Analysis Chemical Application

Alka. CaC03 pH Temp. Turb. Alum

Soda

Alum

Soda

Ash Chlorine

Polymer

8170

Polymer 

8101

Polymer 

NP10

Jan.

Avg.

Max.

Min.

29.60

38.00

23.00

38.50

50.00

30.00

7.65

8.04

6.25

5.96

6.50

5.50

1.24

5.60

0.78

4.05 1.04 11.73 1.64 1.03 12.22

Feb.

Avg.

Max.

Min.

28.33

33.00

24.00

38.21

60.00

32.00

7.58

7.97

7.21

6.01

6.40

5.60

1.47

9.60

0.81

3.74 1.15 14.18 1.65 1.01 12.39

Mar.

Avg.

Max.

Min.

28.70

38.00

14.00

37.00

44.00

23.00

7.60

8.16

7.13

6.30

7.00

5.40

1.41

4.50

0.90

4.47 1.03 14.2 1.51 1.00 12.06

Apr.

Avg.

Max.

Min.

29.72

34.00

16.00

36.48

44.00

32.00

7.53

8.16

6.99

6.88

8.70

6.00

1.79

8.90

0.88

4.59 1.01 14.79 1.32 1.03 12.33

May

Avg.

Max.

Min.

26.02

30.00

22.00

33.31

48.00

28.00

7.59

8.49

7.10

7.50

8.30

6.30

2.66

10.40 

0.89

6.60 1.02 13.18 1.36 1.02 14.75

June

Avg.

Max.

Min.

25.78

30.00

20.00

30.86

40.00

26.00

7.44

8.40

6.65

6.93

12.00 

5.20

2.26

21.00 

0.83

5.86 0.97 13.56 1.30 1.00 15.97

July

Avg.

Max.

Min.

26.00

34.00

24.00

26.00

36.00

24.00

7.36

8.80

6.20

7.48

8.50

6.40

1.72

9.20

0.95

5.81 0.83 14.56 1.33 1.03 18.19

Aug.

Avg.

Max.

Min.

25.69

34.00

24.00

28.53

34.00

23.00

6.93

7.80

6.70

7.78

9.10

7.00

2.18

3.00

1.06

5.53 1.05 16.55 1.33 1.03 20.29

Sept.

Avg.

Max.

Min.

26.30

30.00

24.00

28.40

32.00

25.00

6.97

7.70

6.70

8.63

10.00 

7.50

1.66

4.40

1.15

5.55 1.02 18.01 1.33 1.00 20.41

Oct.

Avg.

Max.

Min.

25.71

30.00

22.00

28.07

34.00

20.00

7.23

7.70

6.60

10.93 

12.90 

9.50

1.61

5.30

0.73

5.75 1.13 13.71 1.41 0.98 21.95

Nov.

Avg.

Max.

Min.

26.23

32.00

21.00

28.57

32.00

26.00

7.51

7.99

6.89

8.80

10.50 

6.20

1.09

3.00

0.73

4.55 0.95 8.58 1.42 1.01 20.85

Dec.

Avg.

Max.

Min.

25.46

42.00

22.00

28.09

40.00

24.00

7.62

7.89

7.19

6.93

8.00

6.00

1.07

5.30

0.75

3.83 0.99 6.58 1.29 0.99 16.22

Note: Polymer 8170 dosage figures are in ppb. All other chemicals are expressed in mg/L.

270
Table D-2. Raw Water, Pine Valley, 1988

Water Analysis Chemical Application

Alka. CaC03 pH Temp. Turb. Alum

Soda

Alum

Soda

Ash Chlorine

Polymer

8170

Polymer 

8101

Polymer 

NP10

Jan.

Avg.

Max.

Min.

24.48

28.00

20.00

26.03

31.00

20.00

7.61

8.16

7.16

6.29

6.80

5.60

1.15

4.90

0.74

3.92 8.96 0.89 1.36 1.02 15.65

Feb.

Avg.

Max.

Min.

24.40

28.00

20.00

27.35

32.00

20.00

7.44

7.97

7.09

6.58

7.60

6.00

1.30

5.60

0.92

3.62 9.80 1.00 1.24 1.02 18.29

Mar.

Avg.

Max.

Min.

24.89

31.00

22.00

29.55

36.00

26.00

7.25

7.54

6.92

6.39

6.90

5.50

1.27

4.70

0.79

3.49 10.23 1.03 1.22 1.00 16.91

Apr.

Avg.

Max.

Min.

24.77

32.00

19.00

29.71

34.00

28.00

7.34

7.65

6.94

6.53

7.60

5.70

1.43

12.30 

0.62

5.42 11.25 0.63 1.20 0.80 15.47

May

Avg.

Max.

Min.

23.47

30.00

20.00

28.75

34.00

24.00

7.52

8.29

7.09

7.51

8.60

6.30

1.40

3.70

0.55

4.19 9.58 0.96 1.26 0.99 20.23

June

Avg.

Max.

Min.

24.53

30.00

22.00

28.38

40.00

24.00

7.43

7.83

7.06

8.88

10.60 

7.50

1.29

8.00

0.66

3.59 10.33 0.96 1.26 1.01 0.57

July

Avg.

Max.

Min.

24.11

32.00

22.00

28.42

34.00

23.00

7.41

7.70

7.14

9.40

12.90 

7.60

1.40

4.40

0.84

3.54 10.36 0.96 1.25 1.01 20.54

Aug.

Avg.

Max.

Min.

24.42

28.00

22.00

28.23

32.00

24.00

7.36

7.62

7.11

9.64

10.50 

9.00

1.16

5.20

0.74

3.55 12.17 0.91 1.28 1.01 19.95

Sept.

Avg.

Max.

Min.

23.80

28.00

18.00

28.32

32.00

26.00

7.34

7.65

7.06

10.19 

11.00 

8.90

 1.51

5.70

0.80

3.44 12.72 1.00 1.16 0.99 20.09

Oct.

Avg.

Max.

Min.

23.85

26.00

21.00

28.69

32.00

26.00

7.51

7.92

7.06

11.70 

13.00 

10.40 

1.74

4.60

0.63

3.52 11.23 0.99 1.12 0.99 19.64

Nov.

Avg.

Max.

Min.

23.19

28.00

20.00

27.79

31.00

26.00

7.85

8.33

7.32

9.27

11.70 

6.50

1.09

3.37

0.59

3.96 7.32 1.01 1.30 0.96 18.68

Dec.

Avg.

Max.

Min.

23.19

26.00

20.00

26.85

32.00

20.00

7.90

8.25

7.51

5.68

6.80

4.30

0.93

2.20

0.59

3.91 6.75 1.04 1.17 14.23 0.99

Note: Polymer 8170 dosage figures are in ppb. All other chemicals are expressed in mg/L.

271
Table D-3. Raw Water, Pine Valley, 1989

Water Analysis Chemical Application

Alka. CaC03 pH Temp. Turb. Alum

Soda

Alum

Soda

Ash Chlorine

Polymer

8170

Polymer 

8101

Polymer

NP101P

Jan.

Avg.

Max.

Min.

22.35

26.00

17.00

27.31

32.00

24.00

7.88

8.63

7.49

4.96

5.30

4.40

0.62

1.88

0.47

3.93 7.53 1.02 1.15 13.78 0.97 13.67

Feb.

Avg.

Max.

Min.

22.34

28.00

16.00

26.96

30.00

26.00

7.89

8.13

7.46

4.95

5.40

44.00 

0.70

2.50

0.54

4.00 7.92 1.02 1.12 12.73 1.00 10.79

Mar.

Avg.

Max.

Min.

21.05

24.00

16.00

26.77

30.00

24.00

7.97

8.41

7.32

5.65

6.40

5.00

0.73

6.00

0.43

4.33 8.18 0.92 1.16 0.28 0.97 0.24

Apr.

Avg.

Max.

Min.

21.90

25.00

17.00

27.00

30.00

25.00

7.89

8.34

7.45

6.52

7.80

5.80

0.65

4.30

0.44

4.69 8.45 0.96 1.15 1.09 16.33

May

Avg.

Max.

Min.

21.48

26.00

17.00

26.13

29.00

22.00

7.87

8.44

7.40

7.84

8.60

6.80

0.85

3.90

0.49

3.88 8.65 0.76 1.18 18.34 1.18 13.16

June

Avg.

Max.

Min.

21.00

73.00

 7.00

26.00

28.00

24.00

7.84

8.45

7.00

8.60

9.30

7.70

1.11

4.50

0.04

3.81 9.68 0.61 1.15 17.32 1.03 17.53

July

Avg.

Max.

Min.

20.00

24.00

12.00

26.00

32.00

18.00

7.80

8.53

7.31

9.40

10.40 

8.70

1.20

4.90

0.86

3.66 11.41 0.98 1.11 19.78 1.06 17.71

(NP10P)

Aug.

Avg.

Max.

Min.

21.00

28.00

15.00

26.00

32.00

21.00

7.77

8.38

7.28

10.00 

10.70 

8.70

1.17

4.70

0.71

3.69 12.66 0.93 1.13 18.28 1.01 0.77

(Polymer

8100)

Sept.

Avg.

Max.

Min.

23.00

27.00

20.00

27.00

24.00

30.00

7.71

8.21

7.25

11.00 

12.20 

10.10 

1.22

5.70

0.81

3.90 13.72 1.00 1.19 18.42 1.01

Oct.

Avg.

Max.

Min.

23.00

27.00

19.00

27.00

32.00

24.00

7.76

8.35

7.24

12.50 

14.20 

11.00 

1.29

12.40 

0.62

5.03 12.50 1.01 1.20 16.42 0.52

(Polymer

8100)

Nov.

Avg.

Max.

Min.

23.00

28.00

18.00

26.00

28.00

24.00

7.84

8.25

7.48

8.80

11.20 

6.70

1.49

15.00 

0.66

6.03 8.53 0.95 1.16 15.97 0.51

(Polymer

8100)

Dec.

Avg.

Max.

Min.

23.00

28.00

19.00

27.00

30.00

24.00

7.94

8.23

7.62

5.80

6.80

4.90

1.08

2.90

0.45

6.38 8.64 1.05 1.11 16.35 0.51

(Polymer

8100)

Note: Polymer 8170 dosage figures are in ppb. All other chemicals are expressed in mg/L.

272
Table D-4. Raw Water, Pine Valley, 1990

Water Analysis Chemical Application

Alka. CaC03 pH Temp. Turb. Alum

Soda

Alum

Soda

Ash Chlorine

Polymer

8100

Polymer 

8170

Polymer 

NP10

Jan.

Avg.

Max.

Min.

23.00

28.00

14.00

27.00

31.00

24.00

7.96

8.30

7.36

5.50

5.80

5.10

1.02

1.98

0.74

5.56 9.30 1.00 1.13 0.52 16.02

Feb.

Avg.

Max.

Min.

21.00

26.00

14.00

27.00

36.00

22.00

7.71

8.04

7.41

5.70

6.10

5.40

0.91

3.40

0.54

5.77 8.36 1.00 1.14 0.52 16.19

Mar.

Avg.

Max.

Min.

22.00

28.00

16.00

27.00

32.00

24.00

7.47

9.99

7.18

5.90

6.20

5.50

0.88

2.30

0.62

5.90 10.40 1.04 1.23 0.51 16.16

Apr.

Avg.

Max.

Min.

22.00

28.00

19.00

26.00

30.00

21.00

7.50

7.90

7.30

6.20

6.80

6.00

0.75

3.00

0.48

6.29 11.75 0.56 1.11 0.51 16.22

May

Avg.

Max.

Min.

21.00

26.00

17.00

25.00

28.00

22.00

7.72

7.91

7.11

7.20

8.20

6.10

0.97

3.80

0.66

6.65 9.24 0.74 1.13 0.51 16.31

June

Avg.

Max.

Min.

20.00

26.00

18.00

25.00

28.00

22.00

7.55

8.54

7.21

8.70

9.40

7.80

1.16

4.40

0.71

6.77 10.49 0.92 1.08 0.57 15.65

July

Avg.

Max.

Min.

20.00

26.00

16.00

25.00

27.00

20.00

7.43

7.61

7.18

9.60

11.00 

8.90

1.58

10.20 

0.96

6.78 11.53 1.04 1.14 0.56 16.42

Aug.

Avg.

Max.

Min.

21.00

25.00

18.00

25.00

28.00

20.00

7.35

7.51

7.21

9.60

10.20 

9.10

1.31

4.90

0.92

6.31 11.95 1.06 1.13 0.50 16.11

Sept.

Avg.

Max.

Min.

21.00

25.00

18.00

25.00

28.00

20.00

7.29

7.90

7.15

9.90

10.50 

9.60

1.03

4.30

0.68

6.08 12.45 1.02 1.13 0.50 16.21

Oct.

Avg.

Max.

Min.

21.00

26.00

18.00

26.00

28.00

20.00

7.26

7.45

7.04

9.90

10.60 

9.60

1.07

9.60

0.77

5.81 14.40 0.96 1.18 0.50 16.11

Nov.

Avg.

Max.

Min.

21.00

26.00

18.00

25.00

34.00

14.00

7.95

8.55

7.14

9.40

10.70 

7.60

0.85

1.81

0.58

6.67 9.30 0.75 1.20 0.53 16.04

Dec.

Avg.

Max.

Min.

21.00

26.00

19.00

25.00

28.00

22.00

8.06

8.75

7.54

5.90

7.80

4.20

0.76

1.06

10.50 

7.00 8.69 0.88 1.15 0.51 16.05

Note: Polymer 8170 dosage figures are in ppb. All other chemicals are expressed in mg/L.

273
Table D-5. Raw Water, Pine Valley, 1991

Water Analysis Chemical Application

Alka. CaC03 pH Temp. Turb. Alum

Soda

Alum

Soda

Ash Chlorine

Polymer

8100

Polymer 

8170

Polymer 

NP10

Jan.

Avg.

Max.

Min.

22.00

28.00

18.00

25.00

28.00

22.00

7.76

8.02

7.22

5.20

5.50

4.80

1.34

54.00 

0.41

7.29 10.13 0.97 1.27 0.50 15.74

Feb.

Avg.

Max.

Min.

21.00

28.00

19.00

26.00

30.00

24.00

7.73

7.97

7.51

5.50

6.20

5.10

0.73

3.50

0.33

7.64 10.40 1.01 1.37 0.49 16.39

Mar.

Avg.

Max.

Min.

21.00

26.00

15.00

26.00

30.00

22.00

7.57

7.81

7.29

5.90

6.30

5.40

0.62

2.50

0.37

7.56 11.11 1.02 1.38 0.49 16.28

Apr.

Avg.

Max.

Min.

21.00

24.00

17.00

25.00

28.00

20.00

7.59

7.90

7.14

6.20

6.70

6.00

0.84

7.20

0.45

7.57 10.34 1.03 1.21 0.50 16.32

May

Avg.

Max.

Min.

20.00

22.00

18.00

24.00

26.00

22.00

7.68

7.94

7.47

7.70

8.90

6.00

0.95

3.70

0.67

7.53 9.67 0.98 1.19 0.50 16.00

June

Avg.

Max.

Min.

20.00

24.00

18.00

24.00

26.00

22.00

7.42

7.66

7.09

8.80

9.60

8.00

1.00

3.10

0.72

7.49 11.35 0.90 1.15 0.51 18.20

July

Avg.

Max.

Min.

21.00

26.00

18.00

24.00

28.00

22.00

9.50

10.60 

8.80

7.34

7.68

7.04

0.92

3.00

0.69

7.50 12.62 0.96 1.10 0.51 18.23

Aug.

Avg.

Max.

Min.

20.00

23.00

17.00

24.00

30.00

20.00

9.40

10.30 

8.90

7.14

7.30

7.04

1.06

2.70

0.70

7.34 14.68 0.99 1.21 0.51 18.27

Sept.

Avg.

Max.

Min.

21.00

24.00

18.00

24.00

30.00

20.00

9.40

10.40 

9.00

7.10

7.25

6.99

1.13

3.70

0.82

7.32 16.14 1.00 1.21 0.49 18.00

Oct.

Avg.

Max.

Min.

21.00

24.00

18.00

24.00

28.00

20.00

9.80

10.60 

9.20

7.00

7.63

6.88

1.11

2.90

0.68

6.76 16.89 1.09 1.20 0.51 18.81

Nov.

Avg.

Max.

Min.

20.00

22.00

17.00

24.00

28.00

20.00

8.50

10.60 

6.40

7.71

7.99

7.29

0.90

46.00 

0.48

7.42 9.63 0.98 1.43 0.51 17.89

Dec.

Avg.

Max.

Min.

20.00

22.00

18.00

24.00

28.00

20.00

5.40

6.60

4.90

7.78

7.95

7.56

0.58

1.61

0.43

7.31 9.67 0.58 1.33 0.49 18.09

Note: Polymer 8170 dosage figures are in ppb. All other chemicals are expressed in mg/L.

274
Table D-6. Raw Water, Pine Valley, 1992

Water Analysis Chemical Application

Alka. CaC03 pH Temp. Turb. Alum

Soda

Alum

Soda

Ash Chlorine

Polymer

8100

Polymer 

8170

Polymer 

NP10

Jan.

Avg.

Max.

Min.

20.00

22.00

18.00

24.00

26.00

20.00

5.20

5.70

4.80

7.60

7.79

7.37

0.47

1.43

0.24

7.11 10.27 1.08 1.32 0.50 18.33

Feb.

Avg.

Max.

Min.

20.00

22.00

16.00

24.00

32.00

20.00

4.90

5.40

4.70

7.39

7.56

7.21

0.72

1.94

0.43

6.95 11.97 0.89 1.44 0.50 17.98

Mar.

Avg.

Max.

Min.

19.00

22.00

15.00

24.00

30.00

22.00

5.20

5.50

4.90

7.37

7.82

7.13

0.88

10.80 

0.53

7.02 12.22 1.03 1.45 0.50 18.20

Apr.

Avg.

Max.

Min.

20.00

24.00

18.00

24.00

30.00

16.00

5.80

6.80

5.20

7.41

8.31

7.04

0.92

2.80

0.57

7.58 12.04 1.07 1.41 0.52 17.95

May

Avg.

Max.

Min.

20.00

24.00

17.00

23.00

26.00

19.00

7.40

8.10

6.60

7.46

7.72

7.29

0.85

3.10

0.48

7.54 11.41 0.85 1.41 0.49 18.83

June

Avg.

Max.

Min.

20.00

27.00

17.00

23.00

24.00

20.00

8.20

8.80

7.40

7.30

7.51

7.15

0.57

2.30

0.32

7.83 12.60 1.04 1.41 0.51 22.29

July

Avg.

Max.

Min.

19.00

24.00

16.00

23.00

28.00

19.00

8.60

9.20

8.20

7.18

7.80

7.02

0.68

2.10

0.43

7.92 14.20 0.99 1.39 0.51 23.80

Aug.

Avg.

Max.

Min.

19.00

22.00

16.00

24.00

30.00

20.00

9.00

10.30 

7.80

7.08

7.51

6.90

0.70

3.70

0.47

7.98 15.80 0.91 1.37 0.51 24.08

Sept.

Avg.

Max.

Min.

20.00

23.00

17.00

24.00

30.00

20.00

8.80

9.20

8.20

7.07

7.23

6.87

0.89

2.20

0.53

7.99 17.07 0.91 1.38 0.49 23.76

Oct.

Avg.

Max.

Min.

20.00

22.00

17.00

24.00

26.00

18.00

9.40

9.90

8.80

7.03

7.25

6.90

1.12

2.60

0.79

7.50 17.24 0.95 1.41 0.50 23.99

Nov.

Avg.

Max.

Min.

20.00

22.00

17.00

23.00

26.00

18.00

8.70

10.30 

6.60

7.74

8.06

6.97

0.75

3.40

0.47

8.17 11.52 0.30 1.43 0.50 23.82

Dec.

Avg.

Max.

Min.

20.00

22.00

17.00

24.00

26.00

20.00

5.50

6.60

5.00

7.97

8.37

7.51

0.64

20.00 

0.39

8.52 10.94 1.44 0.50 24.16

Note: Polymer 8170 dosage figures are in ppb. All other chemicals are expressed in mg/L.

275
Appendix E

Chemical Monthly Average Doses, Mesa Water Treatment Plant, 

Colorado Springs, CO, 1987–1992 (Mesa, 1994)

Table E-1. Raw Water, Mesa, 1992

Water Analysis Chemical Application

Alka. CaC03 pH Temp. Turb. Alum

Soda

Ash

Soda

Alum Chlorine

Polymer

8100

Polymer

8170

Polymer 

NP10

Jan.

Avg.

Max.

Min.

40.14

52.00

22.00

39.88

52.00

20.00

7.65

7.86

7.26

40.55

43.00

39.00

2.26

6.10

1.40

5.52 6.65 3.91 1.49 0.71 8.00

Feb.

Avg.

Max.

Min.

22.71

28.00

19.00

23.17

28.00

18.00

7.35

7.50

7.20

40.65

43.00

49.00

2.85

79.50 

1.60

3.99 11.57 6.36 1.62 0.80 1.20

Mar.

Avg.

Max.

Min.

21.75

26.00

18.00

23.11

28.00

18.00

7.35

8.09

7.11

41.67

43.00

41.00

2.20

7.00

1.40

4.08 10.51 6.21 1.62 0.79

Apr.

Avg.

Max.

Min.

20.14

24.00

16.00

21.21

30.00

16.00

7.42

7.69

7.04

42.66

46.00

41.00

2.88

32.00 

1.40

4.25 11.36 6.09 1.70 0.80 7.20

May

Avg.

Max.

Min.

18.89

24.00

16.00

19.37

25.00

16.00

7.49

7.74

7.17

46.53

50.00

44.00

1.63

15.50 

0.60

4.49 11.37 3.84 1.75 0.60 8.60

June

Avg.

Max.

Min.

18.05

22.00

16.00

17.52

22.00

12.00

7.15

7.69

6.43

50.25

55.00

45.00

2.22

58.10 

0.60

6.13 14.87 3.82 2.03 0.67 8.10

July

Avg.

Max.

Min.

17.82

24.00

14.00

16.63

20.00

12.00

7.29

7.73

7.05

53.33

56.00

51.00

1.74

20.70 

0.80

4.93 11.45 6.30 2.04 0.89 0.80

Aug.

Avg.

Max.

Min.

18.75

24.00

14.00

17.89

28.00

12.00

7.15

7.52

6.83

53.25

58.00

49.00

3.42

75.50 

1.00

5.11 13.12 6.20 2.11 0.97 0.30

Sept.

Avg.

Max.

Min.

18.47

24.00

14.00

17.78

28.00

14.00

7.02

7.35

6.73

51.91

55.00

49.00

1.65

19.50 

0.90

5.33 13.90 6.24 2.06 0.94 1.30

Oct.

Avg.

Max.

Min.

23.42

60.00

16.00

22.07

54.00

16.00

7.35

7.87

7.07

49.53

52.00

47.00

3.45

100.00 

0.70

5.43 11.41 5.76 1.79 0.90 5.30

Nov.

Avg.

Max.

Min.

43.20

60.00

22.00

40.95

56.00

20.00

7.90

8.73

7.56

41.00

48.00

36.00

3.99

76.00 

0.70

5.54 8.19 2.56 1.47 0.66 8.00

Dec.

Avg.

Max.

Min.

49.37

62.00

34.00

48.56

60.00

30.00

7.92

8.82

7.40

38.26

41.00

36.00

2.27

15.50 

0.80

5.27 8.12 2.41 1.43 0.61 6.80

Note: Polymer 8170 dosage figures are in ppb. All other chemicals are expressed in mg/L.

276
Table E-2. Raw Water, Mesa, 1991

Water Analysis Chemical Application

Alka. CaC03 pH Temp. Turb. Alum

Soda

Ash

Soda

Alum Chlorine

Polymer

8100

Polymer

8170

Polymer 

NP10

Jan.

Avg.

Max.

Min.

28.89

68.00

14.00

28.89

66.00

18.00

7.87

8.22

7.49

38.39

39.00

37.00

1.86

27.00 

0.50

 6.17 8.72 1.36 0.39 15.80

Feb.

Avg.

Max.

Min.

51.70

62.00

20.00

53.34

64.00

24.00

7.92

8.16

7.55

39.19

41.00

37.00

2.66

22.00 

1.40

 7.51 10.03 2.01 0.55 18.40

Mar.

Avg.

Max.

Min.

51.76

82.00

26.00

52.68

78.00

28.00

8.00

8.37

7.60

41.14

44.00

39.00

4.20

20.00 

1.20

 7.76 6.51 1.66 0.59 18.90

Apr.

Avg.

Max.

Min.

44.33

56.00

24.00

45.68

58.00

24.00

8.10

8.64

7.49

43.17

46.00

40.00

7.74

196.00 

0.90

 7.96 5.63 1.79 0.62 16.80

May

Avg.

Max.

Min.

38.60

54.00

18.00

39.40

54.00

18.00

7.98

8.62

7.52

47.71

55.00

41.00

6.69

122.30 

0.70

 8.56 6.35 2.05 0.69 17.20

June

Avg.

Max.

Min.

27.34

54.00

14.00

28.15

54.00

14.00

7.74

8.24

7.39

52.41

58.00

47.00

13.33 

230.00 

0.70

10.57 10.43 2.39 1.13 14.60

July

Avg.

Max.

Min.

24.52

44.00

12.00

24.19

40.00

12.00

7.45

7.94

6.93

56.43

60.00

53.00

10.14 

900.00 

0.70

10.28 14.36 2.37 0.86 17.30

Aug.

Avg.

Max.

Min.

26.46

40.00

16.00

25.92

40.00

16.00

7.60

8.04

7.15

55.96

61.00

50.00

13.82 

180.00 

1.10

11.99 14.37 0.14 2.22 1.11 15.90

Sept.

Avg.

Max.

Min.

27.68

48.00

14.00

26.98

46.00

16.00

7.56

7.91

7.08

54.28

59.00

50.00

5.28

370.00 

0.70

 6.48 8.24 1.14 2.11 0.80 14.30

Oct.

Avg.

Max.

Min.

27.48

44.00

16.00

26.03

40.00

14.00

7.64

7.98

7.18

50.07

54.00

42.00

1.67

7.50

0.60

 5.90 7.78 0.97 1.88 0.67 8.80

Nov.

Avg.

Max.

Min.

38.96

58.00

20.00

38.60

56.00

22.00

7.88

8.08

7.58

41.78

45.00

39.00

3.01

31.00 

0.70

 6.01 5.81 1.12 1.65 0.71 11.80

Dec.

Avg.

Max.

Min.

42.16

54.00

26.00

41.96

58.00

30.00

7.83

8.19

7.47

40.23

42.00

38.00

2.50

35.00 

0.60

 5.75 5.54 1.03 1.56 0.66 9.20

Note: Polymer 8170 dosage figures are in ppb. All other chemicals are expressed in mg/L.

277
Table E-3. Raw Water, Mesa, 1990

Water Analysis Chemical Application

Alka. CaC03 pH Temp. Turb. Alum

Soda

Ash

Soda

Alum Chlorine

Polymer

8100

Polymer

8170

Polymer 

NP10

Jan.

Avg.

Max.

Min.

43.70

58.00

40.00

46.50

58.00

34.00

7.77

8.10

7.46

39.90

41.00

39.00

1.73

6.20

1.10

5.75 10.32 1.42 0.38 14.24

Feb.

Avg.

Max.

Min.

47.60

70.00

24.00

48.40

66.00

28.00

7.81

8.16

7.49

40.30

43.00

39.00

2.19

34.30

0.50

5.27 8.81 1.30 0.33 12.60

Mar.

Avg.

Max.

Min.

40.90

60.00

24.00

43.60

62.00

26.00

7.66

8.01

7.32

41.30

42.00

39.00

5.47

107.80

1.20

6.61 12.00 1.56 0.47 16.70

Apr.

Avg.

Max.

Min.

42.50

62.00

20.00

44.10

66.00

22.00

7.78

8.09

7.29

43.70

47.00

41.00

15.49

180.00

1.70

9.40 13.92 1.72 0.76 14.92

May

Avg.

Max.

Min.

31.30

46.00

15.00

33.40

46.00

16.00

7.77

8.04

7.34

46.50

53.00

41.00

7.11

91.70 

0.60

9.72 14.66 1.90 0.79 14.61

June

Avg.

Max.

Min.

30.10

44.00

14.00

29.90

44.00

16.00

7.39

7.92

6.91

55.50

59.00

48.00

7.91

615.00

0.80

5.61 14.55 2.38 1.57 16.88

July

Avg.

Max.

Min.

26.00

46.00

12.00

25.30

42.00

10.00

7.31

8.87

6.82

56.20

62.00

52.00

47.50

2,700.00

1.00

7.79 19.61 2.71 2.00 14.59

Aug.

Avg.

Max.

Min.

24.50

42.00

14.00

23.80

40.00

10.00

7.56

8.05

6.80

54.90

60.00

51.00

11.60

1,200.00

0.70

6.84 14.48 2.39 1.55 18.31

Sept.

Avg.

Max.

Min.

33.10

54.00

12.00

31.80

52.00

15.00

7.52

7.94

7.18

54.80

57.00

51.00

3.62

48.00

0.70

5.19 12.74 2.09 1.41 18.59

Oct.

Avg.

Max.

Min.

30.90

61.00

14.00

30.00

64.00

16.00

7.73

8.13

7.34

46.40

54.00

44.00

3.04

63.60

0.60

5.16 11.92 1.95 1.42 16.54

Nov.

Avg.

Max.

Min.

41.10

96.00

16.00

40.00

84.00

16.00

8.05

8.25

7.83

42.20

46.00

39.00

1.59

51.00

0.50

3.21 10.97 1.61 1.22 15.95

Dec.

Avg.

Max.

Min.

54.50

72.00

20.00

58.80

70.00

24.00

8.01

8.25

7.54

38.60

41.00

39.00

1.16

6.40

0.60

6.65 9.38 1.27 0.62 16.58

Note: Polymer 8170 dosage figures are in ppb. All other chemicals are expressed in mg/L.

278
Table E-4. Raw Water, Mesa, 1989

Water Analysis Chemical Application

Alka. CaC03 pH Temp. Turb. Alum

Soda

Ash

Soda

Alum Chlorine

Polymer

8100

Polymer

8170

Polymer 

NP10

Jan.

Avg.

Max.

Min.

46.50

60.00

24.00

42.30

62.00

26.00

7.86

8.12

7.58

48.40

50.00

47.00

3.51

135.10 

1.30

6.07 8.06 0.97 1.26 0.90 14.19

Feb.

Avg.

Max.

Min.

38.20

50.00

20.00

39.80

65.00

18.00

7.62

8.02

7.11

48.80

50.00

47.00

8.44

133.00 

1.20

7.99 16.92 0.97 1.84 5.90 1.33 8.56

Mar.

Avg.

Max.

Min.

38.70

48.00

38.00

41.50

52.00

24.00

7.79

8.13

7.53

49.30

55.00

43.00

12.66 

288.00 

1.80

9.13 15.95 9.13 1.75 7.45 1.17 8.89

Apr.

Avg.

Max.

Min.

36.38

56.00

16.00

38.90

52.00

16.00

7.90

8.15

7.70

46.70

51.00

43.00

4.93

79.50 

0.80

7.78 10.52 1.63 12.82 0.79 0.82

(N101P)

May

Avg.

Max.

Min.

27.60

42.00

16.00

30.50

44.00

18.00

7.88

8.14

7.61

51.10

56.00

47.00

4.31

73.00 

1.00

7.37 11.35 1.53 0.73 11.30

(N101P)

June

Avg.

Max.

Min.

23.70

46.00

12.00

25.90

50.00

14.00

7.64

7.85

7.26

55.10

58.00

52.00

3.09

43.00 

1.40

8.50 14.83 1.81 0.90 14.28

(N101P)

July

Avg.

Max.

Min.

27.90

44.00

16.00

29.70

46.00

18.00

7.31

7.57

6.90

58.00

60.00

56.00

7.71

56.00 

1.80

9.57 18.69 2.05 11.24 1.10 12.80

(N101P)

Aug.

Avg.

Max.

Min.

30.70

60.00

14.00

30.10

50.00

14.00

7.44

7.85

6.96

56.70

62.00

53.00

11.95 

430.00 

1.00

9.62 17.63 1.93 12.54 1.09

Sept.

Avg.

Max.

Min.

30.00

50.00

18.00

30.80

46.00

20.00

7.34

7.72

7.05

54.00

58.00

50.00

5.06

475.00 

0.80

8.47 18.46 1.73 11.64 0.90

Oct.

Avg.

Max.

Min.

34.70

48.00

18.00

36.60

54.00

22.00

7.89

8.59

7.28

51.00

55.00

44.00

1.64

15.20 

0.50

7.25 11.30 1.71 12.39 0.75 0.48

(Polymer 

8100)

Nov.

Avg.

Max.

Min.

34.60

58.00

14.00

35.60

67.00

20.00

8.06

8.32

7.71

42.60

45.00

40.00

0.69

4.50

0.30

6.07 9.11 1.59 16.53 0.39

(Polymer 

8100)

Dec.

Avg.

Max.

Min.

50.10

62.00

18.00

52.40

64.00

20.00

7.99

8.26

7.55

39.60

42.00

39.00

1.18

7.10

0.40

6.09 7.89 1.50 16.78 0.42

(Polymer 

8100)

Note: Polymer 8170 dosage figures are in ppb. All other chemicals are expressed in mg/L.

279
Table E-5. Raw Water, Mesa, 1988

Water Analysis Chemical Application

Alka. CaC03 pH Temp. Turb. Alum

Soda

Ash

Soda

Alum Chlorine

Polymer

8100

Polymer

8170

Polymer 

NP10

Jan.

Avg.

Max.

Min.

50.90

66.00

40.00

 47.40

 62.00

 38.00

7.94

8.10

7.69

41.20

43.00

39.00

 4.71

 200.00

 1.00

6.38 7.14 0.19 1.22 0.70 12.05

Feb.

Avg.

Max.

Min.

49.10

70.00

20.00

 46.60

 68.00

 22.00

7.91

8.13

7.45

40.40

42.00

30.00

 6.19

 77.00

 1.10

6.13 6.97 1.02 1.43 0.72 12.17

Mar.

Avg.

Max.

Min.

48.50

70.00

22.00

 44.60

 60.00

 20.00

7.77

808.00 

731.00 

38.30

41.00

35.00

 16.21

 340.00

 1.00

6.33 8.60 1.00 1.70 0.96 14.53

Apr.

Avg.

Max.

Min.

32.30

70.00

10.00

 30.00

 68.00

 12.00

7.77

8.33

6.67

39.80

45.00

34.00

 7.26

1,480.00

 1.00

8.48 10.96 1.01 1.59 1.01 15.03

May

Avg.

Max.

Min.

31.50

50.00

14.00

 28.20

 44.00

 8.00

7.87

8.34

7.59

45.80

52.00

41.00

 2.99

 25.00

 1.00

7.84 11.29 0.37 1.69 1.06 18.19

June

Avg.

Max.

Min.

23.30

38.00

8.00

 21.00

 35.00

 8.00

7.71

8.15

6.80

55.50

59.00

48.00

 10.64

 492.00

 1.10

9.89 14.58 1.95 1.27 15.80

July

Avg.

Max.

Min.

26.20

158.00

12.00

 23.70

126.00 

 10.00

7.53

8.35

7.19

57.50

63.00

55.00

 6.51

 567.00

 0.80

9.49 15.40 1.87 1.15 16.83

Aug.

Avg.

Max.

Min.

23.00

46.00

12.00

 20.90

 40.00

 12.00

7.54

8.70

7.19

61.00

63.00

58.00

 9.92

1,330.00

 0.80

9.70 15.96 2.02 1.32 14.55

Sept.

Avg.

Max.

Min.

28.50

42.00

14.00

 23.30

 34.00

 12.00

7.58

8.07

7.21

58.80

62.00

56.00

 3.70

 100.00

 0.80

8.09 13.66 1.72 0.88 14.88

Oct.

Avg.

Max.

Min.

34.30

50.00

20.00

 31.90

 48.00

 17.00

7.67

8.17

7.15

55.70

60.00

63.00

 1.57

 16.10

 0.70

7.20 12.84 1.53 0.85 13.91

Nov.

Avg.

Max.

Min.

34.50

42.00

22.00

 32.00

 41.00

 22.00

7.97

8.25

7.37

51.20

55.00

48.00

 1.30

 525.00

 0.70

7.34 10.52 1.42 0.73 17.42

Dec.

Avg.

Max.

Min.

42.90

60.00

22.00

 38.70

 52.00

 18.00

7.86

8.13

7.44

48.40

49.00

47.00

 6.14

 168.00

 0.60

7.30 10.30 0.34 1.35 0.82 18.64

Note: Polymer 8170 dosage figures are in ppb. All other chemicals are expressed in mg/L.

280
Table E-6. Raw Water, Mesa, 1987

Water Analysis Chemical Application

Alka. CaC03 pH Temp. Turb. Alum

Soda

Ash

Soda

Alum Chlorine

Polymer

8100

Polymer

8170

Polymer 

NP10

Jan.

Avg.

Max.

Min.

35.00

66.00

18.00

35.50

68.00

18.00

7.69

8.19

7.32

36.90

39.00

34.00

 3.18

 47.00 

 0.60

 8.23 13.58 1.51 0.95 21.31

Feb.

Avg.

Max.

Min.

34.70

56.00

18.00

36.50

64.00

18.00

7.64

8.17

7.13

38.80

41.00

37.00

 7.04

 86.00 

 1.60

 9.18 12.41 1.72 1.08 20.12

Mar.

Avg.

Max.

Min.

49.80

66.00

18.00

53.90

72.00

20.00

7.74

8.13

7.17

38.70

46.00

35.00

 63.65 

 925.00 

 2.10

12.91 10.73 2.47 1.68 18.89

Apr.

Avg.

Max.

Min.

38.20

58.00

20.00

43.50

64.00

24.00

7.68

8.11

7.14

43.40

54.00

37.00

 17.45 

 208.00 

 2.30

11.88 14.23 1.77 1.57 13.91

May

Avg.

Max.

Min.

25.20

44.00

16.00

29.40

44.00

20.00

7.58

8.57

6.59

50.10

57.00

42.00

 27.63 

1,554.00 

 1.40

11.31 16.32 1.70 1.34 17.37

June

Avg.

Max.

Min.

21.70

44.00

10.00

24.30

46.00

16.00

7.47

8.17

6.63

59.40

66.00

51.00

 22.03 

1,073.00 

 2.50

12.30 18.89 2.12 1.63 16.89

July

Avg.

Max.

Min.

24.80

40.00

12.00

26.90

46.00

16.00

7.54

8.58

6.81

63.50

67.00

59.00

 6.28

 27.00 

 2.50

10.92 18.41 1.94 1.28 17.78

Aug.

Avg.

Max.

Min.

26.00

50.00

12.00

28.20

46.00

14.00

7.47

8.88

6.73

64.30

68.00

61.00

 12.76 

 533.00 

 0.80

10.58 17.25 2.05 1.41 16.75

Sept.

Avg.

Max.

Min.

38.30

58.00

18.00

40.40

58.00

20.00

7.78

8.24

7.29

61.00

65.00

57.00

 24.25 

 861.00 

 2.90

10.15 10.06 1.73 1.17 15.95

Oct.

Avg.

Max.

Min.

29.40

54.00

14.00

30.20

50.00

15.00

7.85

8.29

7.31

53.90

58.00

50.00

 1.76

 16.60 

 0.40

 7.58 10.96 1.66 0.81 17.92

Nov.

Avg.

Max.

Min.

32.90

68.00

16.00

32.10

54.00

15.00

8.04

8.73

7.59

50.60

56.00

46.00

 1.11

 3.10

 0.30

 6.14 7.99 1.43 0.67 13.73

Dec.

Avg.

Max.

Min.

36.10

58.00

18.00

35.20

56.00

20.00

7.97

8.24

7.31

44.70

48.00

42.00

 1.62

 15.40 

 0.30

 5.75 7.79 1.24 0.58 12.56

Note: Polymer 8170 dosage figures are in ppb. All other chemicals are expressed in mg/L.

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