United States Office of Research and EPA/600/R-09/012
Environmental Protection Development January 2009
Agency Washington, DC 20460
vxEPA Nutrient Control Design Manual
State of Technology
Review Report
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EPA/600/R-09/012
January 2009
Nutrient Control Design Manual
State of Technology Review Report
by
The Cadmus Group, Inc
57 Water Street
Watertown, MA 02472
Scientific, Technical, Research, Engineering, and Modeling Support (STREAMS)
Task Order 68
Contract No. EP-C-05-058
George T. Moore, Task Order Manager
United States Environmental Protection Agency
Office of Research and Development/ National Risk Management Research Laboratory
26 West Martin Luther King Drive, Mail Code 445
Cincinnati, Ohio, 45268
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Notice
This document was prepared by The Cadmus Group, Inc. (Cadmus) under EPA Contract No. EP-C-
05-058, Task Order 68. The Cadmus Team was lead by Patricia Hertzler and Laura Dufresne with Senior
Advisors Clifford Randall, Emeritus Professor of Civil and Environmental Engineering at Virginia Tech and
Director of the Occoquan Watershed Monitoring Program; James Barnard, Global Practice and
Technology Leader at Black & Veatch; David Stensel, Professor of Civil and Environmental Engineering at
the University of Washington; and Jeanette Brown, Executive Director of the Stamford Water Pollution
Control Authority and Adjunct Professor of Environmental Engineering at Manhattan College.
Disclaimer
The views expressed in this document are those of the individual authors and do not necessarily,
reflect the views and policies of the U.S. Environmental Protection Agency (EPA). Mention of trade
names or commercial products does not constitute endorsement or recommendation for use. This
document has been reviewed in accordance with EPA's peer and administrative review policies and
approved for publication.
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Abstract
This EPA document is an interim product in the development of revised design guidance for
nitrogen and phosphorus control at municipal WWTPs. This document presents findings from an
extensive review of nitrogen and phosphorus control technologies and techniques currently applied and
emerging at municipal wastewater treatment plants (WWTP). It includes information on the importance
of nutrient removal, the properties and analytical techniques for nitrogen and phosphorus species, and
the principles behind biological nitrogen and phosphorus removal and chemical phosphorus
precipitation. The report profiles the latest advances in technology to achieve consistently low nutrient
levels in plant effluent, including effluent filtration and advanced clarification techniques, along with up-
to-date research on the removal of emerging microcontaminants such as endocrine-disrupting
compounds. Other contemporary issues include how mathematical modeling can improve process
design, nutrient removal at small and decentralized treatment systems, and sustainable nutrient
recovery.
This report was submitted in fulfillment of EPA Contract No. EP-C-05-058, Task Order 68, by The
Cadmus Group, Inc. under the sponsorship of the United States Environmental Protection Agency. This
report covers a period from November 2007 through September 2008 and work was completed as of
October 2008.
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Contents
1. Introduction 13
1.1 Purpose 13
1.2 Methodology 13
1.3 Organization of the Report 13
2. Need for Nitrogen and Phosphorus Removal at Wastewater Treatment Plants 15
2.1 Status of Wastewater Treatment in the U.S 15
2.2 Nutrient Impairment of U.S. Waterways 16
2.3 Federal and State Initiatives to Reduce Nutrient Pollution 19
2.3.1 NPDES Permitting 19
2.3.2 Water Quality Trading 20
2.3.3 Technology Evaluation and Guidance 20
2.4 Industry Initiatives-The Nutrient Removal Challenge Program 21
2.5 Barriers to Implementing Nutrient Removal 22
3. Nutrient Constituents in Wastewater and Measurement Methods 24
3.1 Nitrogen 24
3.2 Phosphorus 26
4. Phosphorus Removal by Chemical Addition 28
4.1 Principles 28
4.2 Location of Chemical Feed and Mixing 30
4.3 Advanced Solids Separation Processes 31
4.4 Other Design and Operational Issues 32
4.5 Impacts on Sludge Handling and Production 32
5. Biological Nitrogen Removal 33
5.1 Principles 33
5.1.1 Nitrification 33
5.1.2 Denitrification 34
5.2 Current Configurations 35
5.2.1 Biological Nitrogen Removal Process Configurations 35
5.2.1.1 Suspended Growth Systems 35
5.2.1.2 Attached Growth and Hybrid Systems 38
5.2.2 Separate Stage Nitrification and Denitrification Systems 39
5.2.2.1 Suspended Growth Nitrification 39
5.2.2.2 Attached Growth Nitrification 40
5.2.2.3 Separate-Stage Denitrification 41
5.3 Key Design and Operational Issues 42
5.4 Guidance for Selecting Process Modifications 44
5.5 Impacts on Sludge Production and Handling 45
6. Biological Phosphorus Removal and Combination Processes 46
6.1 Principles 46
6.2 Current Configurations 48
6.2.1 Pho-redox (A/O) and 3 Stage Pho-redox (A2/O) 48
6.2.2 Modified Bardenpho 49
6.2.3 University of Cape Town (UCT) and Modified UCT (MUCT) 49
6.2.4 Johannesburg (JHB), Modified Johannesburg and Westbank 50
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6.2.5 Orange Water and Sewer Authority (OWASA) 51
6.2.6 Oxidation Ditches 52
6.2.7 Sequencing Batch Reactors (SBR) 53
6.2.8 Hybrid Chemical / Biological Processes 53
6.3 Emerging Technologies 53
6.4 Operational and Design Considerations 54
6.4.1 COD:P Ratio 54
6.4.2 Retention Time 56
6.4.3 Temperature 57
6.4.4 Presence of Oxygen or Nitrate in the Aerobic Zone 57
6.4.5 Avoiding Backmixing of Oxygen 57
6.4.6 pH 57
6.4.7 Anaerobic Release 58
6.4.8 Sufficient Oxygen in the Aerobic Zone 59
6.4.9 Inhibition 59
6.4.10 Flow and Load Balancing 60
6.5 Impacts on Sludge Handling and Removal 60
6.6 Guidance for Selecting Process Modifications 60
6.7 Ongoing Research 61
7. Effluent Filtration 62
7.1 Types of Filters 62
7.1.1 Conventional Down-flow Filters 62
7.1.2 Deep-bed Down-flow Filters 63
7.1.3 Continuous Backwashing Upflow Sand Filters 63
7.1.4 Pulsed Bed Filters 63
7.1.5 Traveling-Bridge Filters 63
7.1.6 Fuzzy Filters 63
7.1.7 Discfilters 64
7.1.8 Cloth Media Disk Filters 64
7.1.9 Membranes 64
7.1.10 Blue PRO™ Process 64
7.1.11 Pressure Filters 65
7.2 Design and Operating Principles 65
7.3 Ongoing Research and Emerging Technologies 66
8. Mathematical Modeling 67
8.1 The Need for Models 67
8.2 Overview of Available Models 67
8.3 Model Inputs 69
8.4 Model Calibration 69
9. Nutrient Removal for Small Communities and Decentralized Wastewater Treatment Systems.. 70
9.1 Phosphorus Removal 70
9.2 Nitrogen Removal 70
9.3 Nitrogen and Phosphorus Removal Technologies 71
9.3.1 Introduction 71
9.3.2 Nutrient Removal technologies 71
10. Sustainable Nutrient Recovery 77
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11. Co-Removal of Emerging Contaminants 79
11.1 Background on Emerging Contaminants 79
11.2 Removal of Emerging Contaminants by Nutrient Removal Technologies 80
References 90
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Tables
Table 3-1. Research Topics on Dissolved Organic Nitrogen Measurement, Fate, and
Environmental Impacts 26
Table 8-1. Available Activated Sludge Models 68
Table 11-1. Estrogens 80
Table 11-2. Study Design Parameters 85
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Figures
Figure 2-1. Population served by POTWs nationwide 16
Figure 5-1. Modified Ludzck Ettinger Process 36
Figure 5-2. Bardenpho Process (Four-Stage) 36
Figure 5-3. Oxidation Ditch with Aerobic and Anoxic Zones 37
Figure 5-4. Step Feed Biological Nitrogn Removal 38
Figure 6-1. Theory of biological phosphorus removal in activated sludge 47
Figure 6-2. Pho-redox Process (A/O) 49
Figure 6-3. 3 Stage Pho-redox Process (A2/O) 49
Figure 6-4. UCTand Modified UCT Process 50
Figure 6-5. JHB and Modified JHB Process 51
Figures 6-6. Westbank Process 51
Figure 6-7. OWASA Process 52
Figure 6-8. Example of Secondary release in Second Anoxic Zone 58
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Acronyms and Abbreviations
A/O
A2/O
AT3
AOB
ASM
BABE
BAF
BAR
BCFS
bDON
BHRC
BNR
BOD
BOD5
BPR
COD
CWA
CWSRF
CSO
DAF
DO
DON
Ei
E2
EBPR
EDC
EOT A
EE2
EPA
FFS
FWPCA
FWS
GAO
HRSD
HRT
iDON
ISF
IWA
JHB
MAUREEN
MBR
MBBR
MGD
MLE
Pho-redox
3 Stage Pho-redox
Aeration Tank 3
Ammonia Oxidizing Bacteria
Activated Sludge Model
Bio-Augmentation Batch Enhanced
Biological Aerated Filter
Bio-Augmentation Regeneration/Reaeration
Biological Chemical Phosphorus and Nitrogen Removal
Biodegradable Fraction of Dissolved Organic Nitrogen
Ballasted High Rate Clarification Processes
Biological Nutrient Removal
Biochemical Oxygen Demand
Biochemical Oxygen Demand (5-day)
Biological Phosphorus Removal
Chemical Oxygen Demand
Clean Water Act
Clean Water State Revolving Fund
Combined Sewer Overflow
Dissolved Air Flotation
Dissolved Oxygen
Dissolved Organic Nitrogen
Estrone
17 R-estradiol
Enhanced Biological Phosphorus Removal
Endocrine Disrupting Chemicals
Ethylene Diamine Tetraacetic Acid
17ct-ethynylestradiol
U.S. Environmental Protection Agency
Fixed-film Systems
Federal Water Pollution Control Act
Free Water Surface
Glycogen Accumulating Organism
Hampton Roads Sanitation District
Hydraulic Retention Time
Inert Dissolved Organic Nitrogen
Intermittent Sand Filter
International Water Association
Johannesburg
Mainstream Autotrophic Recycle Enhanced N-removal
Membrane Bioreactor
Moving-Bed Biofilm Reactor
Million Gallons per Day
Modified Ludzack Ettinger
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MUCT
N
NOAA
NOB
NPDES
NTT
ORD
ORP
OWASA
OWM
P
PAH
PAD
PHA
PHB
PHV
POTW
PPCPs
RAS
RBC
rbCOD
rDON
RO
RSF
SAV
SBR
SHARON
SND
SRT
SSO
STAC
SWIS
TDS
TKN
TMDL
TN
TP
TSS
TUDP
UCT
USDA
USGS
VIP
VFA
VSS
WAS
WEF
Modified University of Capetown
Nitrogen
National Oceanic and Atmospheric Administration
Nitrite Oxidizing Bacteria
National Pollutant Discharge Elimination System
Nitrogen Trading Tool
EPA Office of Research and Development
Oxidation Reduction Potential
Orange Water and Sewer Authority
EPA Office of Wastewater Management
Phosphorus
Polycyclic Aromatic Hydrocarbons
Phosphate Accumulating Organism
Polyhydroxyalkanoates
Poly-B-hydroxy-butyrate
Poly-hydroxy valerate
Publicly Owned Treatment Works
Pharmaceuticals and Personal Care Products
Return Activated Sludge
Rotating Biological Contactor
Readily Biodegradable Chemical Oxygen Demand
Recalcitrant Dissolved Organic Nitrogen
Reverse Osmosis
Recirculating Sand Filters
Submerged Aquatic Vegetation
Sequencing Batch Reactors
Single Reactor High-activity Ammonia Removal Over Nitrite
Simultaneous Nitrification-Denitrification
Solids Retention Time
Sanitary Sewer Overflow
Chesapeake Bay Program Scientific and Technical Advisory
Committee
Subsurface Wastewater Infiltration System
Total Dissolved Solids
Total Kjeldahl Nitrogen
Total Maximum Daily Loads
Total Nitrogen
Total Phosphorus
Total Suspended Solids
Bio-P Model of the Delft University of Technology
University of Capetown
U.S. Department of Agriculture
U.S. Geological Survey
Virginia Initiative Plant
Volatile Fatty Acids
Volatile Suspended Solids
Waste Activated Sludge
Water Environment Federation
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WERF Water Environment Research Foundation
WQS Water Quality Standard
WWTP Wastewater Treatment Plant
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Acknowledgements
The principle authors of this document, titled "Nutrient Control Design Manual: State of
Technology Review Report/' were:
The Cadmus Group, Inc.
Dr. Clifford Randall, Professor Emeritus of Civil and Environmental Engineering at Virginia Tech and
Director of the Occoquan Watershed Monitoring Program
Dr. James Barnard, Global Practice and Technology Leader at Black & Veatch
Jeanette Brown, Executive Director of the Stamford Water Pollution Control Authority and Adjunct
Professor of Environmental Engineering at Manhattan College
Dr. H. David Stensel, Professor of Civil and Environmental Engineering at the University of Washington
EPA technical reviews of the document were performed by:
EPA Office of Research and Development
Donald Brown
George Moore
Douglas Grosse
Richard Brenner
James Smith
Marc Mills
Dan Murray
EPA Headquarters
Donald Anderson
Phil Zahreddine
James Wheeler
EPA Regions
David Pincumbe, Region 1
Roger Janson, Region 1
Dave Ragsdale, Region 10, Office of Water and Watersheds
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External technical reviews of the document were performed by
Dale E. Kocarek, Ohio Water Environment Association
Y. Jeffrey Yang, USEPA Office of Research and Development
Diagrams for illustration of specific concepts were provided by:
Dr. James Barnard, Black and Veatch
Dr. H. David Stensel, University of Washington
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1. Introduction
1.1 Purpose
This document presents findings from an extensive state-of-the-technology review of nitrogen
and phosphorus control technologies and techniques currently applied and emerging at municipal
wastewater treatment plants (WWTP). It includes a description of technologies and identifies key
design and operational issues. Because the majority of WWTPs in the United States are equipped with
secondary biological treatment, the focus of this report is on process and technology
modifications/additions for nutrient removal at existing WWTPs, rather than on new treatment plant
design. Emerging issues such as nutrient removal through decentralized treatment, sustainable
technologies, and co-removal of emerging contaminants are also discussed.
EPA is publishing this document which is an interim product in the development of revised
design guidance for nitrogen and phosphorus control at municipal WWTPs. While the results of the
state-of-the-technology review and the technical reference list presented herein will be the basis for the
revised design guidance manual, which is scheduled to be published in the Fall of 2009, the Agency is
publishing this document at this time to provide this most current technical information in a timely
manner.
1.2 Methodology
The project team began the report with an in-depth literature review using a variety of sources
including: EPA resources across offices and regions, peer-reviewed journals such as Water Research and
Environmental Science and Technology, meeting presentations and workshops, engineering texts,
international reports, government publications, and internet searches. Industry publications such as
Water Environment & Technology and the Water Environment Research Federation reports were also
reviewed. Wastewater treatment experts on the team provided guidance on the sources for the
literature review including unpublished information on current research projects in key areas. Based on
the results of the in-depth literature review, the team identified key findings and prepared technical
summaries for inclusion in this report.
1.3 Organization of the Report
This report is organized into 11 technical chapters as follows:
• Chapter 2. Need for Nitrogen and Phosphorus Removal at Wastewater Treatment
Plants reviews the status of wastewater treatment in the U.S., the impairment of
waterways by excessive nutrients, government and industry initiatives to reduce
nutrient pollution, and the barriers to implementation of such initiatives.
• Chapter 3. Nutrient Constituents in Wastewater and Measurement Methods
describes the forms of nitrogen and phosphorus found in wastewater and the analytical
techniques used to characterize and measure them.
• Chapter 4. Phosphorus Removal by Chemical Addition discusses the principles behind
chemical precipitation, the types of chemicals used, where they are added in the
process train, and traditional and advanced solids separation techniques. The chapter
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also reviews additional design and operational issues as well as how the choice of
chemical impacts sludge handling.
• Chapter 5. Biological Nitrogen Removal examines the principles behind the process,
current and emerging process configurations, key design and operational issues such as
carbon sources and temperature effects, and potential impacts on sludge handling.
• Chapter 6. Biological Phosphorus Removal and Combination Processes discusses the
principles behind biological treatment to remove phosphorus and treatment
configurations that can remove both phosphorus and nitrogen from wastewater. The
chapter provides descriptions of several processes, provides guidance on how to choose
among them, and reviews operational and design considerations including the COD:P
ratio, retention time, and temperature.
• Chapter 7. Effluent Filtration discusses types of filters that can be added as a tertiary
treatment process to WWTPs and summarizes design and operating principles.
• Chapter 8. Mathematical Modeling explains the need for models in designing nutrient
removal processes and examines available models including their input and calibration
requirements.
• Chapter 9. Nutrient Removal for Small Communities and Decentralized Wastewater
Treatment Systems discusses the latest treatment options for on-site wastewater
treatment systems and clustered development systems.
• Chapter 10. Sustainable Nutrient Recovery highlights efforts to develop low-cost and
low-energy technologies to make nutrient removal more efficient, including urine
separation technology and resource recovery from sludge.
• Chapter 11. Co-removal of Emerging Contaminants discusses how some advanced
technologies to remove nitrogen (N) and phosphorus (P) can achieve the additional
benefits of removing some microcontaminants, including endocrine disrupting
compounds (EDCs) and Pharmaceuticals from wastewater.
The References section at the end of the document provides the full, alphabetized list of
technical references reviewed in the development of this report.
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2. Need for Nitrogen and Phosphorus Removal at Wastewater Treatment
Plants
The purpose of this chapter is to provide an overview of the major factors driving decisions to
enhance nutrient removal at WWTPs. Section 2.1 characterizes the industry based on U.S.
Environmental Protection Agency (EPA) survey information. Section 2.2 describes the negative impacts
of nutrient enrichment, highlighting the history of water quality changes in key regions of the country.
EPA and State initiatives to reduce nutrient pollution from wastewater treatment discharges are
summarized in Section 2.3, followed by a discussion of key industry initiatives in Section 2.4. Lastly,
Section 2.5 highlights several barriers to enhancing nutrient removal at wastewater plants.
2.1 Status of Wastewater Treatment in the U.S.
The 1972 Amendments to the Federal Water Pollution Control Act (FWPCA)(Public Law 92-500),
also known as the Clean Water Act (CWA), established the foundation for wastewater discharge control
in the U.S. The CWA's primary objective is to "restore and maintain the chemical, physical, and
biological integrity of the Nation's waters." The CWA established a program to ensure clean water by
requiring permits that limit the amount of pollutants discharged by all municipal and industrial
dischargers into receiving waters. Discharges are regulated under the National Pollutant Discharge
Elimination System (NPDES) permit program. As of 2004, there were 16,583 municipal wastewater
utilities [also known as Publicly Owned Treatment Works (POTWs)] regulated under the CWA, serving
approximately 75 percent of the Nation's population (U.S. Public Health Service and USEPA, 2008) with
the remaining population served by septic or other onsite systems.
Wastewater treatment has generally been defined as containing one or more of the following
four processes: (1) preliminary, (2) primary, (3) secondary, and (4) advanced - also known as tertiary
treatment. Preliminary treatment consists of grit removal, which removes dense inert particles and
screening to remove rags and other large debris. Primary treatment involves gravity settling tanks to
remove settleable solids, including settleable organic solids. The performance of primary settling tanks
can be enhanced by adding chemicals to capture and flocculate smaller solid particles for removal and to
precipitate phosphorus. Secondary treatment follows primary treatment in most plants and employs
biological processes to remove colloidal and soluble organic matter. Effluent disinfection is usually
included in the definition of secondary treatment.
EPA classifies advanced treatment as "a level of treatment that is more stringent than secondary
or produces a significant reduction in conventional, non-conventional, or toxic pollutants present in the
wastewater" (U.S. Public Health Service and USEPA, 2008). Other technical references subdivide
advanced treatment, using the terms "secondary with nutrient removal" when nitrogen, phosphorus, or
both are removed and "tertiary removal" to refer to additional reduction in solids by filters or
microfilters (Tchobanoglous et al, 2003). Effluent filtration and nutrient removal are the most common
advanced treatment processes.
The CWA requires that all municipal wastewater treatment plant discharges meet a minimum of
secondary treatment. Based on data from the 2004 Clean Watersheds Needs Survey, 16,543 municipal
WWTPs (99.8 percent of plants in the country) meet the minimum secondary wastewater treatment
requirements. Of those that provide at least secondary treatment, approximately 44 percent provide
some kind of advanced treatment (U.S. Public Health Service and USEPA, 2008). Figure 2-1 shows how
secondary and advanced wastewater treatment have been implemented since 1940 and also provides
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projected treatment for 2024. Note that "No Discharge" refers to systems that do not discharge treated
wastewater to the Nation's waterways and dispose of wastewater via methods such as industrial reuse,
irrigation, or evaporation.
100.0%
o § 80.0%
60.0%
40.0%
O D)
20.0%
0.0%
D No Discharge
D Advanced
• Secondary
• Less Than Secondary
• Raw
1972 1978 1982 1988 1992 1996 2000 2004
Year
Projected
Figure 2-1. Population served by POTWs nationwide for select years between 1940 and 2004 and projected to 2024 (if all needs are met),
organized by wastewater treatment type.
Source: U.S. Public Health Service and USEPA Clean Watersheds Needs Surveys 2004 Report to Congress (U.S. Public Health Service and USEPA,
2008).
2.2 Nutrient Impairment of U.S. Waterways
The harmful effects of eutrophication due to excessive nitrogen and phosphorus concentrations
in the aquatic environment have been well documented. Algae and phytoplankton growth can be
accelerated by higher concentrations of nutrients as they can obtain sufficient carbon for growth from
carbon dioxide. In addition to stimulating eutrophication, nitrogen in the form of ammonia can exert a
direct demand on dissolved oxygen (DO) and can be toxic to aquatic life. Even if a treatment plant
converts ammonia to nitrate by a biological nitrification process, the resultant nitrate can stimulate
algae and phytoplankton growth. Phosphorus also contributes to the growth of algae. Either nitrogen
or phosphorus can be the limiting nutrient depending on the characteristics of the receiving water.
Nitrogen is typically limiting in estuarine and marine systems and phosphorus in fresh water systems.
According to the 2007 report Effects of Nutrient Enrichment in the Nation's Estuaries: A Decade
of Change, increased nutrient loadings promote a progression of symptoms beginning with excessive
growth of phytoplankton and macroalgae to the point where grazers cannot control growth (Bricker et
al., 2007). These blooms may be problematic, potentially lasting for months at a time and blocking
sunlight to light-dependent submerged aquatic vegetation (SAV). In addition to increased growth,
changes in naturally occurring ratios of nutrients may also affect which species dominate, potentially
leading to nuisance/toxic algal blooms. These blooms may also lead to other more serious symptoms
that affect biota, such as low DO and loss of SAV. Once water column nutrients have been depleted by
phytoplankton and macroalgae and these blooms die, the bacteria decomposing the algae then
consume oxygen, making it less available to surrounding aerobic aquatic life. Consequently, fish and
invertebrate kills may occur due to hypoxia and anoxia, conditions of low to no DO. Eutrophic conditions
may also cause risks to human health, resulting from consumption of shellfish contaminated with algal
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toxins or direct exposure to waterborne toxins. Eutrophication can also create problems if the water is
used as a source of drinking water. Chemicals used to disinfect drinking water will react with organic
compounds in source water to form disinfection byproducts, which are potential carcinogens and are
regulated by EPA.
Advanced eutrophic conditions can lead to "dead zones" with limited aquatic life, which
describes the hypoxia condition that exists in the Northern Gulf of Mexico. A recent U.S. Geological
Survey (USGS) report titled Differences in Phosphorus and Nitrogen Delivery to the Gulf of Mexico from
the Mississippi River Basin documents the contribution of nitrogen and phosphorus from agricultural
and non-agricultural sources in the Mississippi River basin (Alexander et al., 2008). On June 16, 2008 the
joint federal-state Mississippi River/Gulf of Mexico Watershed Nutrient Task Force released its 2008
Action Plan for Reducing, Mitigating, and Controlling Hypoxia in the Northern Gulf of Mexico and
Improving Water Quality in the Mississippi River Basin, which builds upon its 2001 plan by incorporating
emerging issues, innovative approaches, and the latest science, including findings from EPA's Science
Advisory Board. Improvements include more accountability through an Annual Operating Plan, better
tracking of progress, state and federal nutrient reduction strategies, and a plan to increase awareness of
the problem and implementation of solutions (USEPA, 2008b).
Nutrient pollution has also caused significant problems in the Chesapeake Bay. Elevated levels
of both nitrogen and phosphorus are the main cause of poor water quality and loss of aquatic habitats in
the Bay. Significant algae blooms on the water surface block the sun's rays from reaching underwater
bay grasses. Without sunlight, bay grasses cannot grow and provide critical food and habitat for blue
crabs, waterfowl, and juvenile fish. The Chesapeake Bay Program estimates that 22 percent of the
phosphorus loading and 19 percent of the nitrogen loading in the Bay comes from municipal and
industrial wastewater facilities (Chesapeake Bay Program, 2008).
The first national attention to nutrient contamination occurred in the Great Lakes. In the 1960s
Lake Erie was declared "dead" when excessive nutrients in the Lake fostered excessive algae blooms
that covered beaches and killed off native aquatic species due to oxygen depletion. At that time,
phosphorus was the primary nutrient of concern due to the advent of phosphate detergents and
inorganic fertilizers. With the enactment of the CWA and the Great Lakes Water Quality Agreement in
1972, a concerted effort was undertaken to reduce pollutant loadings, including phosphorus in the Lake.
Although the health of the Lake improved dramatically, in recent years, there has been renewed
attention to the re-emergence of a "dead" zone in Lake Erie, again due to nutrient loadings. Recent
studies by scientists and the National Oceanic and Atmospheric Administration (NOAA) have also
hypothesized a relationship between excessive nutrients in the Lake and the presence of two aquatic
invasive species - the zebra mussel and the quagga mussel (Vanderploeg et al., 2008).
Development and population increases in the Long Island Sound Watershed have resulted in a
significant increase in nitrogen loading to the Sound. The increased nitrogen loads have stimulated
plant growth, increased the amount of organic matter settling to the benthic zone, lowered DO levels,
and changed habitats. The primary concerns in the Sound include hypoxia, the loss of sea grass, and
alterations in the food web. Management efforts are currently underway to reduce nitrogen pollution
by more than half with a focus on upgrading WWTPs with new technologies and removing nitrogen by
reducing polluted run-off through best management practices on farms and suburban areas (Long Island
Sound Study, 2004).
The above represent four examples of impaired large water bodies impacted by nutrient
loadings. There are more than 80 additional estuaries and bays, and thousands of rivers, streams, and
lakes that are also impacted by nutrients in the U.S. In fact, all but one state and two territories have
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CWA section 303(d) listed1 water body impairments for nutrient pollution. Collectively, states have
listed over 10,000 nutrient and nutrient-related impairments.
Climate change may also be a significant influence on the development of future eutrophic
symptoms. According to the report Effects of Nutrient Enrichment in the Nation's Estuaries: A Decade of
Change, the factors associated with climate change that are expected to have the greatest impacts on
coastal eutrophication are:
• Increased temperatures
• Sea level rise
• Changes in precipitation and freshwater runoff
Increased temperatures will have several effects on coastal eutrophication. Most coastal species
are adapted to a specific range of temperatures. Increases in water temperatures may lead to expanded
ranges of undesirable species. Higher temperatures may also lead to increased algal growth and longer
growing seasons, potentially increasing problems associated with excessive algal growth and
nuisance/toxic blooms. Additionally, warmer waters hold less DO, therefore potentially exacerbating
hypoxia. Temperature-related stratification of the water column may also worsen, having a further
negative effect on DO levels.
Climate change models predict increased melting of polar icecaps and changes in precipitation
patterns, leading to sea level rise and changes in water balance and circulation patterns in coastal
systems. Sea level rise will gradually inundate coastal lands, causing increased erosion and sediment
delivery to water bodies, and potentially flooding wetlands. The increased sediment load and
subsequent turbidity increase may cause SAV loss. The positive feedback between increased erosion and
algal growth (as erosion increases, sediment associated nutrients also increase, stimulating growth) may
also increase turbidity. The loss of wetlands, which act as nutrient sinks, will further increase nutrient
delivery to estuaries.
Another report titled Aquatic Ecosystems and Global Climate Change - Potential Impacts on
Inland Freshwater and Coastal Wetland Ecosystems in the United States notes that climate change of the
magnitude projected for the U.S. over the next 100 years will cause significant changes to temperature
regimes and precipitation patterns across the U.S. (Poff et al., 2002). Such alterations in climate pose
serious risks for inland freshwater ecosystems (lakes, streams, rivers, wetlands) and coastal wetlands,
and may adversely affect numerous critical services provided to human populations.
These conclusions indicate climate change is a significant threat to the species composition and
function of aquatic ecosystems in the U.S. However, critical uncertainties exist regarding the manner in
which specific species and whole ecosystems will respond to climate change. These arise both from
uncertainties about how regional climate will change and how complex ecological systems will respond.
Indeed, as climate change alters ecosystem productivity and species composition, many unforeseen
ecological changes are expected that may threaten the goods and services that these systems provide to
humans.
1 Required by Section 303(d) of the CWA, the 303(d) list is a list of state's water bodies that do not meet or are not
expected to meet applicable Water Quality Standards with technology-based controls alone.
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2.3 Federal and State Initiatives to Reduce Nutrient Pollution
2.3.1 NPDES Permitting
Established by the FWPCA Amendment of 1972, EPA's NPDES permit program has been the
primary mechanism for controlling pollution from point sources. Point sources are discrete conveyances
such as pipes or man-made ditches. Individual homes that are connected to a municipal system, use a
septic system, or do not have a surface discharge do not need an NPDES permit; however, POTWs and
other facilities must obtain permits if they discharge directly to surface waters.
NPDES permits for wastewater discharges contain, among other information, effluent limits for
"conventional" pollutants such as biochemical oxygen demand (BOD), total suspended solids (TSS), and
pH as well as limits for specific toxicants including various organic and inorganic chemicals. Permits may
also include effluent limits for "non-conventional" pollutants such as nitrogen and phosphorus. Effluent
limits can be technology-based and/or water-quality based. EPA has established technology-based,
secondary treatment effluent limits for BOD as 5-day biochemical oxygen demand (BOD5), TSS, and pH.
Water-quality based effluent limits are set if the technology-based limits are not sufficient to maintain
the water quality standards (WQS) of the receiving water.
Federal and State regulations related to WQSs and Total Maximum Daily Loads (TMDLs) are
expected to drive down NPDES effluent limits for nitrogen and phosphorus. WQS define the goals for a
water body by designating its uses, setting criteria to protect those uses, and establishing provisions to
protect water bodies from pollutants. Criteria can be narrative or numeric. Regulatory agencies can
adopt nutrient criteria to protect a water body against nutrient over-enrichment and eutrophication
caused by nitrogen and phosphorus. In June 1998, EPA issued a National Strategy for the Development
of Regional Nutrient Criteria. This was followed by publication of recommended nutrient criteria for
most streams and lakes in 2001. In a January 9, 2001 Federal Register notice, EPA recommended that
states and other regulatory agencies develop a nutrient criteria plan to outline their process for
adopting such nutrient criteria (Federal Register, 2001). As of May 2007, only a handful of States and
Territories had adopted nutrient criteria for nitrogen and phosphorus (USEPA, 2007a), although many
have made progress in criteria development. In a memo dated May 25, 2007, EPA encouraged all
regulatory agencies to "...accelerate their efforts and give priority to adopting numeric nutrient
standards or numeric translators for narrative standards for all waters in States and Territories that
contribute nutrient loadings to our waterways" (USEPA, 2007b).
CWA Section 303(d) requires states to develop TMDLs for water bodies on the 303(d) list of
impaired waters. A TMDL is a calculation of the maximum amount of a pollutant a water body can
receive and still meet WQS. TMDLs serve as a tool for implementing WQS. The TMDL targets or
endpoints represent a number where the applicable WQS and designated uses (e.g., such as public
water supply, contact recreation, and the propagation and growth of aquatic life) are achieved and
maintained in the water body of concern. TMDLs identify the level of pollutant control necessary to
meet WQS and support the designated uses of a water body. Once a TMDL is set, the total load is
allocated among all existing sources. The allocation is divided into two portions - a load allocation
representing natural and non-point sources and a waste load allocation representing NPDES permitted
point source discharges. In many regions, water bodies have a poor ability to assimilate nutrients or
water bodies are already impaired from past pollution and the water body cannot handle large loads of
additional nutrients. In these cases, TMDLs may require nutrient permit levels to be even lower than
what might be allowed otherwise by nutrient criteria.
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2.3.2 Water Quality Trading
Water quality trading is a market-based approach to improve and preserve water quality.
Trading can provide greater efficiency in achieving water quality goals by allowing one source to meet its
regulatory obligations by using pollutant reductions created by another source that has lower pollution
control costs. For example, under a water quality trading program, a POTW could comply with discharge
requirements by paying distributed sources to reduce their discharges by a certain amount. The use of
geographically-based trading ratios provides an economic incentive, encouraging action toward the
most cost effective and environmentally beneficial projects.
EPA issued a Water Quality Trading Policy in 2003 to provide guidance to States and Tribes on
how trading can occur under the CWA and its implementing regulations. The policy discusses CWA
requirements that are relevant to water quality trading including: requirements to obtain permits,
antibacksliding provisions, development of WQSs including an antidegradation policy, NPDES permit
regulations, TMDLs and water quality management plans. EPA also developed a number of tools and
guidance documents to assist states, permitted facilities, non-point sources, and stakeholders involved
in the development of trading programs (www.epa.gov/owow/watershed/trading.htm). Recently, the
U.S. Department of Agriculture (USDA) National Resources Conservation Service released a Nitrogen
Trading Tool (NTT) prototype for calculating nitrogen credits based on the Nitrogen Loss and
Environmental Assessment Package Model (Gross et al., 2008).
Water quality trading programs have been successfully implemented in several states and
individual watersheds across the county. For example, nitrogen pollution from point sources into the
Long Island Sound was reduced by nearly 25 percent using an innovative Nitrogen Credit Trading
Program. In Connecticut, the program was implemented among 79 sewage treatment plants in the
state. Through the Nitrogen Credit Exchange, established in 2002, the Connecticut program has a goal of
reducing nitrogen discharges by 58.5 percent by 2014.
A recent American Society of Civil Engineers journal article points out, however, that regulatory
frameworks for water quality trading programs have yet to be adopted by the majority of States.
Barriers to adopting such programs include uncertainty in: (1) the mechanisms for determining
appropriate credits and ratios between point sources and distributed sources; and (2) approaches to
ensure that promised reductions actually occur (Landers, 2008).
2.3.3 Technology Evaluation and Guidance
In addition to regulatory and policy initiatives, EPA helps control nutrients through the
development and dissemination of technical information. For example, EPA's Office of Wastewater
Management (OWM) has developed a number of technology fact sheets on secondary and advanced
biological treatment (USEPA, 1999b; 1999c). OWM has also published several technology reports
including Emerging Technologies for Wastewater Treatment and In-Plant Wet Weather Management
(USEPA, 2008a). This technology guide, published in February 2008, is designed to help municipal
wastewater treatment system owners and operators find information on emerging wastewater
treatment and in-plant wet weather management. OWM is also finalizing a document titled Municipal
Nutrient Removal Technologies Reference Document. Although still in draft, volume 1 of the Technical
Report is designed to provide performance and cost information to wastewater facilities on nutrient
removal.
Recently, EPA Region 10 initiated a project to evaluate municipal WWTPs that have
demonstrated exemplary phosphorus removal through their treatment processes. In April 2007, the
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Region published a report titled, Advanced Wastewater Treatment to Achieve Low Concentration of
Phosphorus (EPA Region 10, 2007).
In 1975, EPA's Office of Research and Development (ORD) published its first technology design
guidance for nitrogen removal: Process Design Manual for Nitrogen Control. The manual was updated in
1993 and focused on biological/mechanical processes that were finding widespread application for
nitrification and nitrogen removal at that time. The development of guidance for phosphorus removal
followed a similar schedule, with ORD publishing the document, Process Design Manual for Phosphorus
Removal, in 1971. In 1976, the manual was updated to include design guidance for phosphorus removal
using mineral addition and lime addition. In 1987, EPA published two technical documents to address
phosphorus control: (1) an update to the 1976 Process Design Manual for Phosphorus Removal, and (2)
a handbook titled, Handbook - Retrofitting POTWsfor Phosphorus Removal in the Chesapeake Bay
Drainage Basin. The primary goal of this project was to obtain and share information about the
technology, performance, and costs of applying advanced wastewater treatment for phosphorus
removal. EPA is currently revising these materials to provide updated, state-of-the technology design
guidance for both nitrogen and phosphorus control at municipal WWTPs. This State of Technology
Report is an interim publication in the development of these design guidance documents.
2.4 Industry Initiatives - The Nutrient Removal Challenge Program
In early 2007, The Water Environment Research Foundation (WERF) created a Nutrient Removal
Challenge program the goals of which are to:
• Identify, assess, and make recommendations for improvement of sustainable wastewater
nutrient removal technologies
• Provide information to help agencies meet various receiving water body requirements and
other wastewater treatment goals (e.g., climate change, sustainability, cost-effectiveness,
reliability)
• Research to inform regulatory decision making and help practitioners comply with
increasingly high levels of nitrogen and phosphorus removal with a focus on improving plant
performance
This multi-year program will be funded for 5 years with WERF and external funds anticipated to
total $8-10 million. This research effort will:
• Promote collaborative efforts and engage stakeholders
• Increase technology understanding, explore Limits of Technology (LOT), and reduce costs
• Provide sound scientific information to support regulators, wastewater treatment plant
owners and operators, and other stakeholders
• Leverage WERF research dollars to maximize program contributions and impacts
As part of the Nutrient Challenge's kick-off activities, a WERF Nutrient Research Stakeholder
Workshop was held on March 7 and 8, 2007 in Baltimore, MD, to further refine the Challenge's research
needs and to seek funding partners and collaborators. The facilitated workshop drew almost 100
participants representing all of the key stakeholder groups in the industry. A total of 25 priority areas
were identified, many of them similar to those identified in a similar workshop conducted by WERF in
2006. Generally, these projects fall within the identified top-priority research areas of:
• Characterization of effluent organic nitrogen
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• Accuracy of analytical measurement techniques for low concentrations of phosphorus
• Alternative carbon sources for denitrification
WERF will also be developing a Nutrient Compendium, a comprehensive, living document that
describes the current knowledge of regulatory and technological nutrient removal issues. The
document will describe the key knowledge areas affecting nutrient removal to very low levels and
identifies knowledge gaps related to nutrient removal. Seven topics have been selected as the top
priority. These are:
• Effluent dissolved organic nitrogen
• Alternative carbon sources
• Regulatory issues for low-level nitrogen and phosphorus
• Operations
• Biological treatment processes for achieving low nitrogen and phosphorus effluent levels
• Low phosphorus concentration measurements
• Tertiary phosphorus removal
2.5 Barriers to Implementing Nutrient Removal
There are a number of barriers that may impact forward progress in implementing nutrient
removal processes and achieving reductions in aquatic ecosystems beyond that which is currently being
achieved, including:
• Costs
• Limitations on physical expansion
• State resources
• Increased carbon footprint
• Advanced operations and control
Nutrient removal comes at a cost to municipal wastewater treatment facilities and their
ratepayers. Although funding from the Clean Water State Revolving Fund (CWSRF) is available, it is not
sufficient to address the myriad of CWA-related infrastructure needs (U.S. Public Health Service and
USEPA, 2008).
A second factor affecting the cost of nutrient removal at wastewater facilities is limitations on
physical expansion of wastewater treatment facilities. Some plants are located in urban areas and do
not have any way to obtain the physical space necessary to expand. Space limitations can severely limit
the type of approaches that can be used to reduce nutrients.
In some cases, States are struggling to find the resources to develop WQSs to address nutrient
criteria. Although EPA has developed numerous tools and guidance documents, further technical
assistance may be needed to ensure effective forward progress in this area.
Two potentially negative environmental impacts of employing advanced technologies to remove
nutrients from wastewater are the increase in the carbon footprint and quantity of biosolids requiring
disposal. The increased carbon footprint will result not only from the nitrogen removal process but also
from the increased energy usage necessary to power the technology needed to achieve the proposed
nutrient reduction levels. For utilities in states such as California where greenhouse gas reduction
requirements are on the horizon, increased emissions due to higher energy usage and nitrogen removal
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are cause for serious concern. Higher energy usage may also translate into increased emissions of
airborne nutrients in the form of increased nitrogen oxide emissions.
Operation of biological processes for nitrogen and phosphorus removal requires advanced
knowledge for successful and consistent removal to low effluent levels. The level of process control is
much greater than for BOD5 and TSS removal. In addition, these processes are susceptible to wet
weather, cold weather, and inhibitory substances entering the plants. Plant influent characteristics vary
considerably from one community to another and must be taken into account in the design and
operation of nutrient removal facilities in terms of its impacts on tank volumes and chemical
requirements.
Despite the challenges associated with nutrient removal, new research and information are
steadily becoming available. EPA and industry initiatives discussed earlier in this chapter will continue to
help disseminate this information to wastewater professionals to provide them with the latest
information on nutrient removal strategies.
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3. Nutrient Constituents in Wastewater and Measurement Methods
This chapter provides an overview of the sources, forms, and measurement methods for
nitrogen and phosphorus in wastewater.
3.1 Nitrogen
Nitrogen is an essential nutrient for plants and animals. Approximately 80 percent of the earth's
atmosphere is composed of nitrogen and it is a key element of proteins and cells. The major
contributors of nitrogen to wastewater are human activities such as food preparation, showering, and
waste excretion. The per capita contribution of nitrogen in domestic wastewater is about l/5th of that
for BOD. Total nitrogen in domestic wastewater typically ranges from 20 to 70 mg/L for low to high
strength wastewater (Tchobanoglous et al., 2003). Factors affecting concentration include the extent of
infiltration and the presence of industries. Influent concentration varies during the day and can vary
significantly during rainfall events, as a result of inflow and infiltration to the collection system.
The most common forms of nitrogen in wastewater are:
• Ammonia (NH3)
• Ammonium ion (NH4+)
• Nitrite (NO2~)
• Nitrate (NO3~)
• Organic nitrogen
Nitrogen in domestic wastewater consists of approximately 60 to 70 percent ammonia-nitrogen
and 30 to 40 percent organic nitrogen (Tchobanoglous et al., 2003; Crites and Tchobanoglous, 1998).
Most of the ammonia-nitrogen is derived from urea, which breaks down rapidly to ammonia in
wastewater influent.
EPA approved methods for measuring ammonia, nitrate, and nitrite concentration use
colorimetric techniques. Organic nitrogen is approximated using the standard method for Total Kjeldahl
Nitrogen (TKN) (APHA, AWWA, and WEF, 1998). The TKN method has three major steps: (1) digestion to
convert organic nitrogen to ammonium sulfate; (2) conversion of ammonium sulfate into condensed
ammonia gas through addition of a strong base and boiling; and (3) measurement using colorimetric or
titration methods. Because the measured concentration includes ammonia, the ammonia-nitrogen
concentration is subtracted from the TKN to determine organic nitrogen. Nitrogen components in
wastewater are typically reported on an "as nitrogen" basis so that the total nitrogen concentration can
be accounted for as the influent nitrogen components are converted to other nitrogen compounds in
wastewater treatment.
WWTPs designed for nitrification and denitrification can remove 80 to 95 percent of inorganic
nitrogen, but the removal of organic nitrogen is typically much less efficient (Pehlivanoglu-Mantas and
Sedlak, 2006). Domestic wastewater organic nitrogen may be present in particulate, colloidal or
dissolved forms and consist of proteins, amino acids, aliphatic N compounds, refractory natural
compounds in drinking water (e.g. humic substances), or synthetic compounds (e.g. ethylene diamine
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tetraacetic acid (EDTA)). Organic nitrogen may be released in secondary treatment by microorganisms
either through metabolism or upon death and lysis. Some nitrogen may be contained in recondensation
products. Hydrolysis of particulate and colloidal material by microorganisms releases some organic
nitrogen as dissolved, biodegradable compounds. Amino acids are readily degraded during secondary
biological treatment, with 90 to 98 percent removal in activated sludge systems and 76 to 96 percent
removal in trickling filters. However, other forms of organic nitrogen may be more persistent in
wastewater treatment processes.
The importance of organic nitrogen has increased as effluent limits on nitrogen have become
more stringent. With more impaired waterways from nutrient loads, effluent limits for total nitrogen
(TN) concentrations of 3.0 mg/L or less are becoming more common. The dissolved organic nitrogen
(DON) concentration in the effluent from biological nutrient removal treatment facilities was found to
range from 0.50 to 1.50 mg/L in 80 percent of 188 plants reported by Pagilla (STAC-WERF, 2007) and
values as high as 2.5 mg/L were observed. Thus, for systems without effluent filtration or membrane
bioreactors (MBRs) that are trying to meet a TN treatment goal of 3.0 mg/L, the effluent DON
contribution can easily be 20 to 50 percent of the total effluent nitrogen concentration, compared to
only about 10 percent for conventional treatment (Pehlivanoglu-Mantas and Sedlak, 2004).
The chemical composition of DON in wastewater effluents is not completely understood. Sedlak
(2007) has suggested that only about 20 percent of the DON has been identified as free and combined
amino acids, EDTA, and other trace nitrogen compounds. About 45 percent may be unidentified low
molecular weight compounds and the other 35 percent as unidentified high molecular weight
compounds containing humic acids and amides. Similar results were found by Khan (2007). Early work
by Parkin and McCarty (1981) suggested that 40 to 60 percent of effluent DON is non-bioavailable. The
non-bioavailable portion is also referred to as recalcitrant DON (rDON).
To address these and other water quality issues associated with DON, WERF has identified DON
as a high priority research area. In September 2007, the Chesapeake Bay Program Scientific and
Technical Advisory Committee (STAC) and WERF co-hosted a workshop titled "Establishing a Research
Agenda for Assessing the Bioavailability of Wastewater-Derived Organic Nitrogen in Treatment Systems
and Receiving Waters." The event brought together leaders in the field to discuss the composition,
bioavailability, measurement, and removal of biodegradable DON (bDON) in wastewater treatment and
the composition and test protocol methods for inert DON (iDON) in receiving waters. The WERF
Nutrient Challenge Program has identified key research needs on DON and is fostering collaborative
efforts to make research more efficient and effective. An overview of key areas and researchers
currently working on projects on DON is summarized in Table 3-1.
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Table 3-1. Research Topics on Dissolved Organic Nitrogen Measurement, Fate, and
Environmental Impacts3
Inert DON test
methods
D. Sedlak
UC Berkeley
(WERF)
J. Ma.kinia and K.
Czerwionka
Gdansk University
(Poland Gov't)
Biodegradable DON
and fate in
wastewater
treatment
E. Khan
North Dakota State
University.
(DC WASA)
R. Sharp
Manhattan College
(City of Stamford)
K. Jones
Howard University
(HDR)
K. Pagilla
Illinois Institute of
Technology
(Hazen and Sawyer)
J. Ma.kinia and K. Czerwionka
Gdansk University
(Poland Gov't)
Inert DON
Characteristics
D. Sedlak
UC Berkeley
(WERF)
D. Bronkand E. Canuel
College of William and Mary
M.Mulholland
Old Dominion University
N. Love
University of Michigan
(NSF)
Environmental gradient and
DON
D. Bronkand E. Canuel
College of William and Mary
M. Mulholland
Old Dominion University
N. Love
University of Michigan
(NSF)
a. The sponsoring agency is shown in parentheses
3.2 Phosphorus
Total phosphorus (TP) in domestic wastewater typically ranges between 4 and 8 mg/L but can be
higher depending on industrial sources, water conservation, or whether a detergent ban is in place.
Sources of phosphorus are varied. Some phosphorus is present in all biological material, as it is an
essential nutrient and part of a cell's energy cycle. Phosphorus is used in fertilizers, detergents, and
cleaning agents and is present in human and animal waste.
Phosphorus in wastewater is in one of three forms:
• Phosphate (also called Orthophosphate)
• Polyphosphate, or
• Organically bound phosphorus.
The orthophosphate fraction is soluble and can be in one of several forms (e.g., phosphoric acid,
phosphate ion) depending on the solution pH. Polyphosphates are high-energy, condensed phosphates
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such as pyrophosphate and trimetaphosphate. They are also soluble but will not be precipitated out of
wastewater by metal salts or lime. They can be converted to phosphate through hydrolysis, which is
very slow, or by biological activity (see Chapter 6 for a detailed discussion of biodegradation of
phosphate and polyphosphate in wastewater).
Organically bound phosphorus can either be in the form of soluble colloids or particulate. It can
also be divided into biodegradable and non-biodegradable fractions. Particulate organically bound
phosphorus is generally precipitated out and removed with the sludge. Soluble organically bound
biodegradable phosphorus can be hydrolyzed into orthophosphate during the treatment process.
Soluble organically bound non-biodegradable phosphorus will pass through a wastewater treatment
plant. A typical wastewater contains 3 to 4 mg/L phosphorus as phosphate, 2 to 3 mg/L as
polyphosphate, and 1 mg/L as organically bound phosphorus (WEF and ASCE, 2006).
Phosphorus content in wastewater can be measured as
• Orthophosphate
• Dissolved orthophosphate
• Total phosphorus
• Total dissolved phosphorus (i.e., all forms except particulate organic phosphorus)
EPA approved laboratory methods rely on colorimetric analysis. Colorimetric analysis measures
orthophosphate only, so a digestion step is needed to convert polyphosphate and organic phosphorus
to orthophosphate to measure TP. The persulfate method is reported to be the most common and
easiest method (WEF and ASCE, 2006). To determine dissolved phosphorus (either total dissolved
phosphorus or total dissolved orthophosphate), the sample is first filtered through a 0.45 micron filter.
USEPA approved colorimetric methods are routinely used to measure phosphorus levels as low as 0.01
mg/L. On-line analyzers that use the colorimetric method are available from venders (e.g., the Hach
Phosphax™ SC phosphate analyzer).
Ion chromatography is a second common technique used to measure orthophosphate in
wastewater. As with colorimetric methods, digestion is required for TP measurement, with persulfate
digestion recommended (WEF and ASCE, 2006).
At a workshop in May 2006, WERF members identified a need to develop a standard method for
measurement of very low phosphorus levels in wastewater and characterization of the residual
phosphorus fraction (Bott, 2007). The WERF Nutrient Challenge Research Plan for 2007/2008 identified
"Low Phosphorus Analytical Measurement Reliability" as research to be funded in 2007 or 2008 (WERF,
2007).
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4. Phosphorus Removal by Chemical Addition
The purpose of this chapter is to describe techniques for phosphorus removal by chemical
addition. It summarizes issues associated with chemical feed location, mixing, and sludge production.
An overview of advanced solids separation processes is also provided.
4.1 Principles
Chemical precipitation for phosphorus removal is a reliable, time-tested, wastewater treatment
method that has not drastically changed over the years. To achieve removal, various coagulant aids are
added to wastewater where they react with soluble phosphates to form precipitates. The precipitates
are removed using a solids separation process, most commonly settling (clarification). Chemical
precipitation is typically accomplished using either lime or a metal salt such as aluminum sulfate (alum)
or ferric chloride. The addition of polymers and other substances can further enhance floe formation
and solids settling. Operators can use existing secondary clarifiers or retrofit primary clarifiers for their
specific purposes.
Aluminum and Iron Salts
Alum and ferric or ferrous salts are commonly used as coagulant and settling aids in both the
water and wastewater industry. They are less corrosive, create less sludge, and are more popular with
operators compared to lime. Alum is available in liquid or dry form, can be stored on site in steel or mild
concrete, and has a near unlimited shelf life. Ferric chloride is similar although care is needed during
handling because of corrosivity. If an industrial source is available such as waste pickle liquor, ferrous
chloride or ferrous sulfate have been used for phosphorus removal. Ferrous forms should be added
directly to aerobic reactors rather than to anaerobic reactors such as primary settling basins because the
ferrous iron needs to oxidize to ferric iron for best results.
The following are sample reactions of aluminum and iron salts with phosphorus:
AI2(SO4)3»(14H2O) + 2H2PO4" + 4HCO3" -^ 2AIPO4 + 4CO2 + 3SO4 2" + 18H2O (4-1)
FeCI3»(6H2O) + H2PO4" + 2HCO3" -^ FePO4 + 3CI" + 2CO2 + 8H2O (4-2)
The molar ratio of aluminum to phosphorus required for phosphorus removal ranges from
about 1.38:1 for 75 percent removal, 1.72:1 for 85 percent removal, and 2.3:1 for 95 percent removal.
For iron compounds, a ratio of about 1:1 is required, with a supplemental amount of iron (10 mg/L)
added to satisfy the formation of hydroxide (WEF and ASCE, 1998). For additional removal of
phosphorus with aluminum and iron salts, a ratio of between 2 and 6 parts metal salt to 1 part
phosphorus may be required for adequate phosphorus removal.
To supplement stoichiometry calculations, designers should consider jar tests and, in some
cases, full-scale pilot tests to gauge the effects on the required dose of competing reactions; the
influence of pH and alkalinity, adsorption, and co-precipitation reactions; and the interaction with
polymers that are added to increase coagulation and flocculation (WEF and ASCE, 1998; Bott et al.
2007).
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Aluminum or ferric iron salts can be added to the primary clarifier, secondary clarifier, tertiary
clarifier, or directly into the activated sludge aeration tank. Multiple additions can increase phosphorus
removal efficiency. Ferrous salts can only be added to the aeration basin since it needs to be oxidized to
ferric to precipitate the phosphorus.
The solubility of aluminum and iron salts is a function of pH. The optimum solubility for alum
was previously reported to occur at a pH range of 5.5 to 6.5, significantly lower than most influent
wastewater. Recent studies (Szabo et al., 2008) showed that the range for both iron and alum is
between 3.5 and 7.5 with the highest efficiency between pH 5.5 and 7.
Chemicals such as lime compounds, caustic soda, and soda ash can be used to raise the pH of
the waste stream prior to biological treatment processes or discharge. It is important to understand that
alkalinity is consumed during the precipitation reactions, and precipitation will be incomplete if
insufficient alkalinity is present.
Lime
Although lime had lost favor due to issues associated with chemical handling, sludge production,
and re-carbonation, it has recently been considered more often because of its ability to reduce
phosphorus to very low levels when combined with effluent filtration and the microbial control
properties associated with its high pH. When lime is added to wastewater, it first reacts with the
bicarbonate alkalinity to form calcium carbonate (CaCO3). As the pH increases to more than 10, excess
calcium ions will react with phosphate to precipitate hydroxylapatite [CA5(OH)(PO4)3]. Because it reacts
first with alkalinity, the lime dose is essentially independent of the influent phosphorus concentration.
Tchobanoglous et al. (2003) estimates the lime dose to typically be 1.4 to 1.6 times the total alkalinity
expressed as CaCO3.
The typical reaction between calcium compounds and phosphorus is represented below:
5Ca2+ + 4OH" + 3HPO4. ^ Ca5OH(PO4)3 + 3H2O (4-3)
The molar ratio required for phosphorus precipitation with lime is approximately 5:3, but can
vary from between 1.3 to 2, depending on the composition of the wastewater. As with iron and
aluminum salts, jar tests can be used to determine correct doses for a specific wastewater stream (WEF,
1998).
Lime addition can raise the pH to greater than 11. Because activated sludge processes require
pH levels below 9, lime cannot be added directly to biological treatment processes or it will cause
process upsets. Lime can be added to primary sedimentation tanks and removed with the primary
sludge or it can be added as a tertiary treatment process after biological treatment. When added to
primary tanks, it will also result in the removal of colloidal material through coagulation and settling,
with a concomitant removal of TSS up to 80 percent and chemical oxygen demand (COD) up to 60
percent. In either case, pH adjustment is needed and typically accomplished by adding CO2 or a liquid
acid such as sulfuric acid, nitric acid, or hypochlorite (Tchobanoglous et al., 2003; USEPA, 1999a).
Hortskotte et al. (1974) showed that when the primary effluent is discharged directly to a nitrifying
activated sludge plant, the hydrogen ions produced may neutralize the high pH. However, when
denitrification is practiced and the operator wishes to make use of the soluble COD in the primary
effluent, the effluent must be neutralized before discharging it to the anoxic zone.
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Lime requires special handling and operations practices that further set it apart from chemical
precipitation by metal salts. Although the formation of carbonate scaling on equipment and pipes is a
drawback of lime treatment, lime slaking, where quicklime (CaO) is reacted with water to form calcium
hydroxide (Ca(OH)2), is the biggest operational disadvantage.
Performance Data
Performance data are available for chemical phosphorus removal at many plants, including
Breckenridge, CO, Dillon, CO, Parker, CO, Hillsboro, OR, Tigard, OR, (alum); Walton, NY (aluminum
chloride); Milford, MA (poly-aluminum chloride); Alexandria, VA (ferric chloride and alum); Upper
Occoquan Sewage Authority Plant Centreville, VA (high lime); and the Norman Cole Lower Potomac
Plant, Fairfax County, VA (ferric chloride) (EPA Region 10, 2007).
4.2 Location of Chemical Feed and Mixing
Lime or metal salts can be added at several locations throughout the treatment plant to remove
phosphorus. "Pre-precipitation" is when chemicals are added to raw water to precipitate phosphorus in
the primary sedimentation basins. "Co-precipitation" involves adding chemicals to form precipitates
that can be removed with biological sludge. "Post-precipitation" is when chemicals are added after
secondary sedimentation and precipitants are removed in a tertiary process such as sedimentation or
filtration (Tchobanoglous et al., 2003). Because it requires a high pH to achieve a low phosphorus
concentration, lime cannot be added directly to biological reactors or to the secondary clarifiers. Multi-
point additions of iron or aluminum salts have been very effective and can typically remove more
phosphorus than single-point applications.
There are several advantages to post-precipitating phosphorous using a tertiary treatment
technique (after biological processes in a separate reactor):
• Microorganisms rely on phosphorus as a food source. If too much phosphorus is removed
prior to biological treatment, biological processes may suffer. For activated sludge, the
minimum ratio of phosphorus to BOD5 for a rapidly growing (low solids retention time (SRT))
system is typically about 1:100 (WEF and ASCE, 1998).
• Competing chemicals in the primary sedimentation basins can increase the required dose.
• Phosphorus enters the treatment plant as soluble orthophosphate, soluble polyphosphates,
and organically bound phosphorus. Most of the polyphosphates and much of the
organically bound phosphorus are converted to more simple orthophosphates during
biological treatment. If the influent contains significant polyphosphates and/or organically
bound phosphorus, locating chemical treatment after biological processes would be more
efficient and achieve lower effluent levels.
• The removal of carbonate alkalinity and phosphorus by lime prior to biological treatment
can have a negative impact on nitrification processes (WEF and ASCE, 1998). Also, removing
phosphorus to very low concentrations upstream of denitrification filters can negatively
affect the denitrification process. Previous studies showed that the hydroxide alkalinity can
be balanced by the hydrogen ions produced during nitrification.
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• Sludge recalcification can be used to achieve high removal efficiencies using lime in tertiary
treatment.
One potential advantage to adding chemicals during primary treatment instead of tertiary
treatment is reduced capital costs and space requirements as a result of removing additional BOD and
TSS and reducing the load to downstream processes, thereby reducing the size of the subsequent
activated sludge basins and the amount of oxygen transfer needed.
Chemicals should be well mixed with the wastewater to ensure reaction with soluble
phosphates and formation of precipitates. Chemicals may either be mixed in separate tanks or can be
added at a point in the process where mixing already occurs. Bench-scale and pilot scale tests are often
used to determine the correct mixing rate for a given composition of wastewater and chemicals used,
including polymer (USEPA, 1999a).
4.3 Advanced Solids Separation Processes
The effectiveness of phosphorus removal by chemical addition is highly dependent on the solids
separation process following chemical precipitation. Direct addition of metal salts to activated sludge
processes followed by conventional clarification can typically remove TP to effluent levels between 0.5
and 1.0 mg/L (Bott et al., 2007). Tertiary processes (post-secondary treatment) can be used to remove
phosphorus to very low (< 0.1 mg/L) concentrations. For example, Reardon (2005) reports that four
WWTP with tertiary clarifiers achieved TP levels of between 0.032 and 0.62 mg/L.
Two common tertiary processes are clarification and effluent filtration. These approaches can
be used separately or in combination. Chapter 7 presents a detailed discussion of effluent filtration.
Advances in tertiary clarification processes are discussed below.
The types of clarifiers used for tertiary processes include conventional, one or two-stage lime,
solids-contact, high-rate, and ballasted high-rate (BHRC). Several patented BHRC using different types
of ballast such as recycled sludge, microsand, and magnetic ballast (USEPA, 2008a) have been developed
in recent years. The advantages of high-rate clarification are that the clarifiers have a smaller footprint
and are able to treat larger quantities of wastewater in a shorter period of time. In addition, as an add-
on during wet weather, they can help prevent sanitary sewer overflows (SSOs) and combined sewer
overflows (CSOs). The following patented processes are examples of high rate clarification including
performance estimates:
• DensaDeg® uses a coagulant in a rapid mix basin to destabilize suspended solids. The water
flows into a second tank where polymer (for aiding flocculation) and sludge are added. The
sludge acts as the "seed" for formation of high density floe. This floe is removed in settling
tubes (USEPA, 2008). The main advantages of this process are a smaller footprint and
denser sludge which is easier to dewater. Pilot testing for City of Fort Worth, Texas found a
phosphorus removal rate of 88-95% for DensaDeg® (USEPA, 2003).
• Actiflo® uses a coagulant in a rapid mix basin to destabilize suspended solids. The water
flows to a second tank where polymer (for aiding flocculation) and microsand are added.
Microsand provides a large surface onto which suspended solids attach, creating a dense
floe that settles out quickly. Clarification is assisted by lamella settling. Product pilot testing
in Fort Worth, Texas showed a phosphorus removal efficiency of 92-96% for Actiflo®
(USEPA, 2003).
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• The CoMag process uses the addition of magnetic ballast with metal salts to promote floe
formation. Settling is followed by high gradient magnetic separation for effluent polishing
and recovery of the magnetic ballast (USEPA, 2008a). CoMag is currently in operation at a
4.0 million gallons per day (MGD) wastewater treatment plant in Concord, Massachusetts.
The vendor has guaranteed an effluent phosphorus concentration not to exceed 0.05 mg/L
(EPA Region 10, 2007).
4.4 Other Design and Operational Issues
Phosphorus removal by chemical addition is limited to the soluble phosphates in the waste
stream. Organically bound phosphorus and polyphosphates will not be removed by chemical treatment
unless they are coagulated with the chemicals and removed in the sludge. As noted in Section 4.2,
chemicals can be added after biological treatment to capitalize on the conversion of polyphosphates and
organically bound phosphorus to phosphates by microorganisms in activated sludge.
The success of phosphorus removal by chemical addition depends on proper instrumentation
and control. Dosage control typically takes the form of manual operation (for small systems),
adjustments based on automatic flow measurements, or the more advanced on-line analyzers with
computer-assisted dosage control.
Chemical properties of any water used for making solutions should be considered - tap water
high in suspended solids could cause sludge to form when mixed with coagulants (WEF and ASCE, 1998)
and could lead to clogging of chemical feed lines. Smith et al. (2008) found that factors such as pH,
complexation, mixing, and the coagulant used can limit the removal of phosphorus, especially in the
range of <0.1 mg/L.
4.5 Impacts on Sludge Handling and Production
Sludge handling and production is generally considered to be one of the main downsides of
chemical addition. Chemical precipitation methods always produce additional solids due to generation
of metal- or calcium- phosphate precipitates and additional suspended solids (WEF and ASCE, 1998).
Chemically treated sludge has a higher inorganic content compared to primary and activated sludge and
will increase the required size of aerobic and anaerobic digesters. Additional sludge production can be
estimated using reaction equations.
The use of metal salts can result in increased inorganic salts (salinity) in the sludge and in the
effluent. Salinity can create problems when biosolids are land applied or when the effluent is returned
to existing water supply reservoirs. Biological phosphorus removal was developed in South Africa due to
the high rate of indirect recycling of wastewater effluent which led to excessive total dissolved solids
(TDS) during dry periods. High total salts can reduce germination rates for crops and negatively affect
the soil structure.
Lime traditionally produces a higher sludge volume compared to metal salts because of its
reaction with natural alkalinity. An advantage of lime sludge is that some stabilization can occur due to
the high pH levels required. One disadvantage is that lime can cause scaling in mechanical thickening
and dewatering systems. There are also differences in the amount and characteristics of sludge
generated by alum versus ferric salts. Although alum tends to produce less sludge than do ferric salts,
alum sludge can be more difficult to concentrate and dewater compared to ferric sludge.
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5. Biological Nitrogen Removal
This chapter provides an overview of the principles behind biological nitrogen removal and
describes the common design configurations in use today. It identifies key operational and design issues
(including impacts on sludge handling and production), provides general guidelines on process selection,
and summarizes ongoing research efforts in this area. Process configurations that are designed to
remove both nitrogen and phosphorus are described in Chapter 6.
5.1 Principles
In wastewater treatment, nitrogen removal occurs in two sequential processes: nitrification and
denitrification. An overview of each process is provided below.
5.1.1 Nitrification
Nitrification is an aerobic process in which autotrophic bacteria oxidize ammonia or nitrite for
energy production. Nitrification is normally a two-step aerobic biological process for the oxidation of
ammonia to nitrate. Ammonia-nitrogen (NH3-N) is first converted to nitrite (NO2~) by ammonia oxidizing
bacteria (AOB). The nitrite produced is then converted to nitrate (NO3~) by nitrite oxidizing bacteria
(NOB). Both reactions usually occur in the same process unit at a wastewater treatment plant (e.g.,
activated sludge mixed liquor or fixed film biofilm).
The group of AOB most associated with nitrification is the Nitrosomonas genus, although other
AOB such as Nitrosococcus and Nitrosospira can contribute to the process. Nitrobacter are the NOB
most associated with the second step, although other bacteria including Nitrospina, Nitrococcus, and
Nitrospira have been found to also oxidize nitrite (Tchobanoglous et al., 2003; USEPA, 2007c). AOB and
NOB are classified as autotrophic bacteria because they derive energy from the oxidation of reduced
inorganic compounds (in this case, nitrogenous compounds) and use inorganic carbon (CO2) as a food
source. Nitrifying bacteria require a significant amount of oxygen to complete the reactions, produce a
small amount of biomass, and cause destruction of alkalinity through the consumption of carbon dioxide
and production of hydrogen ions. For each gram (g) of NH3-N converted to nitrate, 4.57 g of oxygen are
used, 0.16 g of new cells are formed, 7.14 g of alkalinity are removed, and 0.08 g of inorganic carbon are
utilized in formation of new cells (Tchobanoglous et al., 2003).
Nitrifying bacteria grow slower and have much lower yields as a function of substrate
consumed, compared to the heterotrophic bacteria in biological treatment processes. The maximum
specific growth rate of the nitrifying bacteria is 10 to 20 times less than the maximum specific growth
rate of heterotrophic bacteria responsible for oxidation of carbonaceous organic compounds in
wastewater treatment. Thus, the nitrification process needs a significantly higher SRTto work
compared to conventional activated sludge processes. The SRT needed for nitrification in an activated
sludge process is a function of the maximum specific growth rate (which is related to temperature), the
reactor dissolved oxygen concentration, and pH. Nitrification rates decline as the DO concentration
decreases below 3.0 mg/L and the pH decreases below 7.0 mg/L. With sufficient DO and adequate pH,
typical nitrification design SRTs range from 10 to 20 days at 10°C and 4 to 7 days at 20°C (Randall et al.,
1992).
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5.1.2 Denitrification
In municipal and industrial wastewater treatment processes, denitrification is the biological
reduction of nitrate or nitrite to nitrogen gas (N2) as indicated by equation 5-1 below.
NOs -» N02 -» NO -» N20 -» N2 (5-1)
It is accomplished by a variety of common heterotrophic microorganisms that are normally present in
aerobic biological processes. Most are facultative aerobic bacteria with the ability to use elemental
oxygen, nitrate, or nitrite as their terminal electron acceptors for the oxidation of organic material.
Heterotrophic bacteria capable of denitrification include the following genera: Achromobacter,
Acinetobacter, Agrobacterium, Alcaligenes, Arthrobacter, Bacillus, Chromobacterium, Corynebacterium,
Flavobacterium, Hypomicrobium, Moraxella, Nesseria, Paracoccus, Propionibacteria, Pseudomonas,
Rhizobium, Rhodopseudmonas, Spirillum and Vibrio (Tchobanoglous et al., 2003). Recent research has
shown that nitrite reduction is accomplished by a much more specialized group of heterotrophic
bacteria than those performing the conversion of nitrate to nitrite (Katehis, 2007).
Denitrification by heterotrophic nitrifying bacteria and by autotrophic bacteria has also been
observed. An example of a heterotrophic nitrifying bacteria that can denitrify is Parococcus pantotropha,
which obtains energy by nitrate or nitrite reduction while oxidizing ammonia under aerobic conditions.
A readily available carbon source, such as acetate, is needed (Robertson and Kuenen, 1990). The
conditions required for this form of denitrification are not practical in biological wastewater treatment.
An autotrophic denitrifying bacteria of practical significance in wastewater treatment is that in the
Anammox process used to remove nitrogen in return streams from anaerobic digestion sludge
dewatering filtrate or centrate. These bacteria have been identified as a member of bacteria in the order
Planctomycetales (Strous et al, 1999). Under anaerobic conditions, ammonia is oxidized with the
reduction of nitrite with the final product as nitrogen gas. The reaction is best accomplished at
temperatures above 25 C and they are slow growing organisms.
Facultative denitrifying bacteria will preferentially use oxygen instead of nitrate. In the absence
of oxygen, however, they will carry out nitrite and/or nitrate reduction. Microbiologists generally use
the term anaerobic to describe biological reactions in the absence of oxygen. To distinguish anaerobic
conditions for which the biological activity occurs mainly with nitrate or nitrite as the electron acceptor,
the term "anoxic" has been applied.
Although oxygen is known to inhibit denitrification, denitrification has been observed in
activated sludge and fixed film systems in which the bulk liquid DO concentration is positive. This is due
to the establishment of an anoxic zone within the floe or biofilm depth. Hence, a single system can carry
out simultaneous nitrification and denitrification. The DO concentration that is possible for simultaneous
nitrification and denitrification depends on a number of factors including the mixed liquor
concentration, temperature, and substrate loading. The DO concentration above which denitrification is
inhibited may vary from 0.10 to 0.50 mg/L (WEF and ASCE, 2006; Tchobanoglous et al., 2003; Barker and
Dold, 1997).
The organic carbon source for denitrifying bacteria can be in the form of:
• Soluble degradable organics in the influent wastewater
• Soluble organic material produced by hydrolysis of influent particulate material
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• Organic matter released during biomass endogenous decay
A general rule of thumb is that 4 g of wastewater influent BOD is needed per g of NO3-N to be
removed through biological treatment (Tchobanoglous et al., 2003). When denitrification occurs after
secondary treatment, there is little BOD remaining so a supplemental carbon source is often needed.
The most common exogenous carbon source in use is methanol; however, due to issues regarding its
safety, cost, and availability, some wastewater systems are using alternative carbon sources such as
acetic acid, ethanol, sugar, glycerol, and proprietary solutions depending on the needs of their particular
facility (deBarbadillo et al., 2008). See Section 5.3 for additional discussion on supplemental carbon
sources.
Biological denitrification reactions produce alkalinity and heterotrophic biomass. Based on the
stoichiometry of the reactions, denitrification will produce a 3.57 mg/L of alkalinity as CaCO3 for each
mg/L of NO3 -N consumed. Heterotrophic biomass produced can be estimated as 0.4 g volatile
suspended solids (VSS) produced for every gram of COD consumed. Growth kinetics for denitrifiers are
dependent on a number of factors including carbon substrate type and concentration, DO
concentration, alkalinity, pH, and temperature, with carbon source being the most important.
5.2 Current Configurations
Biological nitrogen removal can be accomplished by a variety of treatment configurations using
suspended growth, attached growth, or combined systems. In the past, some WWTPs were required to
only remove ammonia-nitrogen in wastewater to reduce toxicity to aquatic organisms with no limits on
nitrate or total nitrogen. However, most treatment plants are now required to remove nitrogen
because both ammonia-nitrogen and nitrate-nitrogen can stimulate algae and phytoplankton growth
and lead to eutrophication of U.S. waterways (See Chapter 2 for additional discussion). For biological
nitrogen removal, it is essential that nitrification occur first followed by denitrification.
Section 5.2.1 presents process configurations that are designed to achieve nitrification and
denitrification in a single process unit. Section 5.2.2 discusses separate-stage nitrification and
denitrification designs.
5.2.1 Biological Nitrogen Removal Process Configurations
Biological nitrogen removal systems achieve nitrification and denitrification along with BOD
reduction in bioreactors followed by secondary clarification. Processes can be either suspended growth
or hybrid systems that use a combination of attached growth (biofilms) and suspended growth
technologies. Configurations within each of these classifications are discussed below. Note that
biological processes that removal both nitrogen and phosphorus are discussed in Chapter 6.
5.2.1.1 Suspended Growth Systems
Modified Ludzck Ettinger (MLE) process
The most common nitrogen removal process used at WWTPs is the Modified Ludzck Ettinger
(MLE) process, which is considered a pre-denitrification, single sludge system. The process includes an
initial anoxic zone, followed by an aerobic zone. In the anoxic zone, nitrate produced in the aerobic zone
is reduced to nitrogen gas. This process uses some of the BOD in the incoming waste. Nitrification
occurs in the aerobic zone along with the removal of most of the remaining BOD. At the end of the
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aerobic zone, pumps recycle the nitrate-rich mixed liquor to the anoxic zone for denitrification. Total
nitrogen removal for the MLE process is typically 80 percent, and the process achieves total effluent
nitrogen concentrations ranging from approximately 5 to 8 mg/L with internal nitrate recycle ratios of 2
to 4 based on the influent flowrate (2-4Q).
Internal Recycle
Influent
RAS
Figure 5-1 Modified Ludzck Ettinger Process
WAS
Four-Stage Bardenpho Process
The four-stage Bardenpho process builds on the MLE process, with the first two stages being
identical to the MLE system (anoxic zone followed by an aeration zone with a nitrate-rich recycle from
the aeration to the anoxic zone). The third stage is a secondary anoxic zone to provide denitrification to
the portion of the flow that is not recycled to the primary anoxic zone. Methanol or another carbon
source can be added to this zone to enhance denitrification. The fourth and final zone is a re-aeration
zone that serves to strip any nitrogen gas and increase the DO concentration before clarification. Some
configurations have used an oxidation ditch instead of the first two stages. This process can achieve
effluent TN levels of 3 to 5 mg/L.
Influent
Internal Recycle
I
RAS
WAS
j Anoxic
] Aerobic
Figure 5-2 Four stage Bardenpho Process
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Sequencing Batch Reactors
Sequencing batch reactors (SBRs) are fill and draw batch systems in which all treatment steps
are performed in sequence for a discreet volume of water in a single or set of reactor basins. SBRs use
four basic phases for most systems:
Fill: water is added to the basin and is aerated and mixed
React: Biological processes are performed
Settle: All aeration and mixing is turned off and the biomass is allowed to settle
Decant: Clarified effluent is removed and biomass is wasted as necessary
The SBR control system allows it to mimic most other suspended growth processes such as the
MLE or Four-Stage Bardenpho system. It typically completes 4 to 6 cycles per day per tank for domestic
wastewater. If properly designed and operated, SBRs can achieve about 90 percent removal of nitrogen
(WEF and ASCE, 2006). SBR applications for small systems are discussed in Section 9.3 of this report.
Oxidation Ditches
Oxidation ditches are looped channels that provide continuous circulation of wastewater and
biomass. A number of operating methods and designs have been developed to achieve nitrogen
removal, all of which work by cycling the flow within the ditch between aerobic and anoxic conditions.
DO can be added to the aerated zone using horizontal brush aerators, diffused aerators with
submersible mixers, or vertical shaft aerators (WEF and ASCE, 2006). Patented designs include the
NITROX process, Carrousel, and BioDenitro (WERF, 2000a). Many oxidation ditch configurations can
achieve simultaneous nitrogen and phosphorus removal. See Section 6.2.6 of this report for additional
information.
ANOXIC ZONE
DO PROBE
WAS
Figure 5-3 Oxidation Ditch with Aerobic and Anoxic Zones
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Step Feed
The step feed biological nitrogen removal process splits the influent flow and directs a portion of
it to each of several anoxic zones, with the highest proportion of influent flow going to the first zone and
steadily decreasing until the last anoxic zone prior to clarification. The biomass in the later stages are
not just treating influent flow but are also used to reduce nitrate from the upstream zones.
The step feed system provides flexibility for systems to handle wet-weather events. It can also
be compatible with existing conventional "plug flow" activated sludge processes and it does not require
the installation of recycle pumps and piping. Disadvantages include the need to control the DO
concentration of aeration zones preceding the downstream anoxic zones and the need to control the
flow splitting to the step feed points.
INFL.
RAS
Anoxic
Aerobic
Figure 5-4 Step Feed Biological Nitrogen Removal
WAS
5.2.1.2 Attached Growth and Hybrid Systems
Integrated Fixed-Film Activated Sludge (IFAS)
Integrated fixed-film activated sludge (IFAS) is any suspended growth system (e.g., MLE, Four-
Stage Bardenpho) that incorporates an attached growth media within the suspended growth reactor in
order to increase the amount of biomass in the basin. IFAS systems have higher treatment rates than
suspended growth systems and generate sludge with better settling characteristics. Many types of fixed
and floating media are available, including:
• Rope: also called looped-cord or strand media. Consists of a polyvinyl chloride-based
material woven into rope with loops along the length to provide surface area for the
biomass (WERF, 2000b). Proprietary designs include Ringlace, Bioweb, and Biomatrix
(USEPA, 2008a).
• Moving Bed with Sponge: proprietary products include Captor and Linpor (USEPA, 2008a).
• Plastic Media: several types of free-floating plastic media are available from Kaldness.
Other media types include fabric mesh (e.g., AccuWeb) and textile material (Cleartec).
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Moving-Bed Biofilm Reactor (MBBR)
The moving-bed biofilm reactor (MBBR) is similar to the IFAS system in that it uses plastic media
with a large surface area to increase biomass within the biological reactor. The MBBR media is
submerged in a completely mixed anoxic or aerobic zone. The plastic media are typically shaped like
small cylinders to maximize surface area for biomass growth. The difference between MBBR and IFAS is
that MBBR does not incorporate return sludge (WERF, 2000b).
Membrane Bioreactor (MBR)
MBRs are commonly designed for nitrogen removal, using membranes for liquid-solids
separation following the anoxic and aerobic zones instead of conventional clarification. Membranes can
be submersed in the biological reactor or located in a separate stage or compartment. Low-pressure
membranes (ultrafiltration or microfiltration) are commonly used. Systems can be pressure driven or
vacuum. All systems use an air scour technique to reduce buildup on the membranes (USEPA, 2007d;
USEPA, 2008a).
Membrane materials are either organic polymers or inorganic materials such as ceramics. They
are designed in modular units and are typically configured as either hollow fiber bundles or plate
membranes (USEPA, 2007d)
For biological nutrient removal applications, the design SRTs and design principals for MBR
systems are similar to those used for systems with secondary clarifiers. One of the main differences is
that the MBR systems operate at a higher MLSS concentration which results in smaller tanks and smaller
space requirements. In addition, membrane separation provides for greatly reduced TSS in the effluent,
typically below 1.0 mg/L, and hence slightly greater removal of nitrogen and phosphorus. Operational
issues include potential for membrane biofouling and increase pumping costs (USEPA, 2007d; WEF,
2005).
5.2.2 Separate Stage Nitrification and Denitrification Systems
5.2.2.1 Suspended Growth Nitrification
Single-sludge systems for BOD removal and nitrification require that the biomass inventory be
retained long enough to establish a stable population of nitrifiers and that the HRT be such that the
biomass can react with the ammonia-nitrogen entering the system. The overall approach for designing
such systems is to determine the target SRT for the system based on influent characteristics (i.e., BOD,
ammonia-nitrogen, organic nitrogen), environmental conditions such as temperature and flow
characteristics (i.e., average daily, maximum monthly, diurnal peak).
Most activated sludge treatment plants will readily nitrify if they have sufficient aerobic SRT and
can deliver sufficient oxygen maintaining 2 mg/L DO or greater. For plants having difficulty in nitrifying
due to insufficient tank volume, there are some emerging technologies which can improve the process.
One of these is bioaugmentation. Bioaugmentation is accomplished by seeding the activated sludge
process with an external source of nitrifying bacteria (also known as external bioaugmentation) or
making process improvements to increase the activity of or enrich the nitrifier population (also known
as in situ bioaugmentation).
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External bioaugmentation uses either commercial sources of nitrifiers or sidestream processes
to grow nitrifiers onsite. Early experiences with commercial sources were not consistent, so most work
to date has been with sidestream production onsite (USEPA, 2008a). Two patented sidestream
configurations for external bioaugmentation are the Single reactor High-activity Ammonia Removal Over
Nitrite (SHARON) process and the In-Nitri® process. Both provide high temperature sidestream
nitrification using ammonia from the anaerobically digested sludge dewatering liquid or digested
supernatant. The nitrifiers grown in the sidestream reactor are fed to the main liquid treatment stream.
Both use flow through reactors with hydraulic retention times (HRT) in the 2 to 3 day range. In the
SHARON process, nitrification is stopped mainly at nitrite by such process control methods as low DO
concentration, low pH and/or low SRT. Full-scale operating systems for the SHARON process include
installations at Utrecht, Rotterdam, Zwolle, Beverwijk, Groningen, The Hague in the Netherlands, and a
system is expected to start up in New York City in 2009. Seeding from a diffused air biological nutrient
removal process to stimulate nitrification in a parallel oxygen process has proved successful at a number
of locations (Bott et al., 2007).
Emerging in situ bioaugmentation technologies used to enhance nitrifier growth and shown to
be successful in bench, pilot, and/or full-scale trials are described briefly below (USEPA, 2008a):
• The Bio-Augmentation Regeneration/Reaeration (BAR) process was developed in the U.S. and is
identical to the Regeneration-DeNitrification (R-DN) process developed independently in the
Czech Republic. It works by recycling ammonia-laden filtrate or centrate from dewatering of
aerobically digested sludge to the head of the aeration tank. The sidestream is fully nitrified,
seeding the aeration tank with additional nitrifying bacteria which allows for reduced SRT.
There are numerous full scale applications in the Czech Republic, U.S., and Canada. The
Aeration Tank 3 (AT3) is similar to the BAR process except that it sends a smaller fraction of the
return activated sludge (RAS) to the aeration tank in order to stop the nitrification process at the
nitrite stage.
• Bio-Augmentation Batch Enhanced (BABE) process uses a SBRto grow nitrifiers by feeding it RAS
and reject water from the sludge dewatering process. After treatment, concentrated nitrifiers
are recycled to the head of the aeration tank.
• The Mainstream Autotrophic Recycle Enhanced N-removal (MAUREEN) Process was developed
for the two-sludge treatment configuration at the Blue Plains Advanced Wastewater Treatment
Plant in Washington, DC. The process involves sidestream treatment of WAS from the second
stage to preferentially select AOB for bioaugmentation to the first sludge stage.
5.2.2.2 Attached Growth Nitrification
Attached growth processes will also nitrify. Trickling filters and rotating biological contactors
(RBCs) have historically been used for biological treatment of wastewater and can achieve nitrification
with a low organic loading and a relatively high media volume. Typically, nitrification is achieved on the
media after most of the BOD is removed since the heterotrophic population competes with the nitrifying
organisms for oxygen and space on the media. A major disadvantage of these technologies compared
to suspended growth systems is that denitrification is fully dependent on addition of a supplemental
carbon source. Suspended growth processes, on the other hand, can be designed to denitrify 80
percent or more of nitrate using the incoming BOD as the carbon source, which is a lower cost solution.
Consequently, trickling filters and RBCs have fallen out of favor for nutrient removal applications.
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In recent years, manufacturers have developed new technologies called biological aerated filters
(BAF) to achieve BOD removal and nitrification. USEPA (2008a) identifies two existing BAF designs as
established technologies: the Biofor® system and the Biostyr® system. The Biofor® filtration system is a
fixed bed, upflow system with a dense granular media that is designed to expand during filtration. Air is
sprayed into the filter to maintain an aerobic environment. The Biostyr® system is similar but uses a
media that is less dense than water and held in place during operation by a screen at the top of the cell.
BAF can be configured in series to remove BOD in one unit and ammonia-nitrogen in the next or
it can be designed for BOD removal and nitrification in a single unit depending on process goals.
Advantages of BAF include its smaller footprint, higher hydraulic loading rates, and less susceptibility to
washout than suspended sludge systems (Verma et al., 2006).
Another fixed film process that has gained popularity lately is moving bed biofilm reactors
(MBBR). These reactors involve biofilm attached to a plastic media in a series of fluidized bed reactors.
The plastic media help promote specialization of the biofilm within each reactor for either nitrification
or denitrification (WEF and ASCE, 2006). Mixers or medium bubble diffuse aeration are used to keep the
media suspended, depending on whether the system is anaerobic or aerobic. MBBR has a shorter SRT
and smaller footprint than activated sludge processes. It has also proven to be effective in cold
temperatures (Bott et al., 2007).
5.2.2.3 Separate-Stage Denitrification
A separate-stage denitrification system may be appropriate for plants that are regularly
achieving nitrification and need to add denitrification capabilities. Attached growth systems
(denitrifying filters) are more common than suspended growth systems, although suspended growth
systems have been used for some treatment plants. Suspended growth reactors typically have short
SRTs (2 to 3 hrs) and a small aerated zone following the denitrification zone to oxidize excess methanol
and release contained nitrogen gas bubbles (WEF and ASCE, 2006).
Denitrification filters typically have a small footprint compared to suspended growth systems
and have the added advantage of achieving denitrification and solids removal simultaneously. They
were first installed in the 1970s and have evolved into two main process configurations (USEPA, 2007c):
• Downflow denitrification filters are deep bed filters consisting of media, support gravel, and
a block underdrain system. Wastewater flow is directed over weirs onto the top of the filter
where a supplemental carbon source, typically methanol, is added. Backwashing (typically
air scouring and backwashing with air and water) is conducted at regular intervals to remove
entrapped solids from the filter. During operation, nitrate is converted to nitrogen gas and
becomes entrained in the filter media, increasing head loss through the filter. To release
entrained nitrogen, most denitrification systems have a nitrogen-release cycle operation
that essentially "bumps" the filter by turning on the backwash pump(s) for a short period of
time.
• Upflow continuous backflow filters do not have to be taken off-line for backwashing, as it is
an integral part of the filtering process. Wastewater enters the bottom of the filter where a
carbon source, typically methanol, is added. Water flows up through an influent pipe and is
dispersed into the filter media through distributors. Filtered water discharges at the top of
the filter. Filter media continuously travels downward, is drawn into an airlift pipe at the
center of the filter, and is scoured before being returned to the filter bed.
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Performance of denitrifying filters depends on many factors including:
• Influent weir configuration
• Filter media
• Underdrain system
• Backwash system
• Flow and methanol feed control
One wastewater system in Connecticut reported that key design issues for them were influent
piping design to minimize aeration, maintaining a consistent flow to the filters, and control of methanol
feed based on influent COD (Pearson et al., 2008).
5.3 Key Design and Operational Issues
Temperature
In general, as temperature of the wastewater increases, the rate of nitrification and
denitrification increases. For the typical range of liquid temperatures between 10 and 25 °C, the
nitrification rate will approximately double for every 8 to 10 °C increase in temperature (WEF and ASCE,
2006). Rapid decreases in temperature without acclimation time will, however, cause even slower
nitrification rates than predicted, strictly by the temperature change. Denitrification rates will also
increase with increasing temperature, although not at the same magnitude as nitrification rates.
Dissolved Oxygen
Nitrifying bacteria are also more sensitive to DO levels as compared to aerobic heterotrophic
bacteria, with growth rates starting to decline below 3 to 4 mg/L with significant reduction below 2
mg/L The nitrification rate at a DO concentration of 0.50 mg/L is only about 60 percent of that at a 2.0
mg/L DO concentration. Studies have shown that the amount of oxygen available to nitrifying bacteria
can be limited by floe size, requiring higher bulk DO concentrations under higher organic loading
conditions (Stenstrom and Song, 1991). At DO concentrations less than 0.5 mg/L, the effect is greater
for Nitrobacterthan for Nitrosomonas. This can result in higher NO2-N in the effluent and have a
negative impact on chlorine disinfection as 1 g of NO2-N consumes 5 g chlorine for oxidation. As noted
in Section 5.1, DO must normally be less than 0.2 to 0.5 mg/L, otherwise there will be inhibition of the
denitrification process.
pH and Alkalinity
Nitrification generally operates well within a pH range of 6.8 to 8.0 (WEF and ASCE, 2006). At
lower pH values the nitrification rate is much slower and at pH values near 6.0 the nitrification rate may
only be about 20 percent of that with a pH of 7.0 (Tchobanoglous et al., 2003). Alkalinity is consumed
during the nitrification process but partially replenished (up to 62.5 percent) during the denitrification
process. Depending on the influent wastewater alkalinity, there may be a sufficient alkalinity reduction
due to nitrification to decrease to unacceptable levels. Addition of chemicals such as lime, sodium
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hydroxide, or soda ash can be used to replace the alkalinity consumed by nitrification to maintain
acceptable pH levels.
Carbon Sources for Denitrification
Denitrifying bacteria need a readily available carbon food source, such as soluble BOD, to
ultimately convert nitrate to nitrogen gas. WWTPs that meet very low total nitrogen limits typically use
a secondary anoxic zone in which supplemental carbon is added. Supplemental sources can be
"internal" such as fermented wastewater or sludge, or "external" sources such as purchased chemicals.
Methanol is currently the most common external carbon source used in denitrification because
of it low cost. It has several drawbacks, however, namely:
• It is highly flammable and implicated in some storage tank explosions and fires (Dolan,
2007); however with proper design and operation problems can be minimized.
• It is not the most efficient source for most treatment configurations.
• Costs have begun to fluctuate widely (deBarbadillo et al., 2008).
• Availability is a problem in some areas (Neethling et al. 2008).
• Reported low growth rates under cold temperatures (Dold et al. 2008).
Other sources of carbon include ethanol, acetic acid, corn syrup, molasses, glucose, glycerol, and
industrial waste products. The WEF Nutrient Challenge Research Plan (2007) identified research on
alternative carbon sources as priority for operators, owners, and engineers of wastewater systems. In
December of 2007, the 2nd External Carbon Workshop was held in Washington, DC to discuss the state of
the technology and research needs. WERF is also currently formulating a standard protocol for
evaluation of external carbon alternatives.
Nitrification Inhibition from Toxic Chemicals
Nitrifying bacteria are very sensitive to heavy metals and other inorganic compounds in
wastewater. The Local Limits Development Guidance Manual (USEPA 2004) has been the main source
of information on inhibitory effects for a variety of wastewater treatment processes including
nitrification. Appendix G of the 2004 version provides a summary table with the reported range of
nitrification inhibition threshold levels for a number of metals, non-metal inorganics, and organic
compounds. Actual inhibitory effects are site-specific and depend on many factors including the nature
of biodegradable organic material, microorganism speciation, acclimation effects, temperature, and
water quality conditions.
Wet Weather Events
Wet weather events can increase inflow and infiltration into the collection system and
subsequently increase the hydraulic load to the wastewater treatment plant. This can in turn reduce the
SRT leading to reduced performance of nitrification process units. In addition, wet weather flows have
different characteristics than typical wastewater influent flow and can be less favorable for nitrification
and denitrification. Conditions that are less favorable for nitrification include decreased alkalinity and
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sudden temperature drops. Lower biodegradable COD concentrations and increased DO make wet
weather flows less amenable to denitrification.
Flow equalization basins can be used to handle wet weather events; however, this requires
available space and capital investment. USEPA (2008a) identifies a number of innovative storage and
treatment technologies used to manage influent flows during wet weather events.
5.4 Guidance for Selecting Process Modifications
Nitrogen removal requires first that a biological nitrification process be present or that the
facility be modified to accomplish nitrification. Considerably more volume is needed for activated
sludge nitrification compared to designs for BOD removal only. If there is insufficient space to
accommodate the increased volume, suspended growth or hybrid process options that require less
space such as the MBR process or IFAS systems with suspended media in the activated sludge process
should be considered. Another option is to use a fixed film nitrification process after the suspended
growth process clarification step. This could be a BAF or a plastic media trickling filter. However, if
nitrogen removal is required, an exogenous carbon source is needed, which has higher operating costs
than using the influent BOD for denitrification.
Nitrification systems need sufficient oxygen transfer for ammonia oxidation in addition to BOD
removal. Such systems should consider the impact to diurnal loadings and ammonia addition in recycle
streams. The influent TN concentration may have daily peak values that are 1.5 to 2.0 times the daily
average loading. Higher peak loadings require longer SRTs to assure that sufficient nitrifying bacteria
are present to remove ammonia at a greater rate, while maintaining a low effluent ammonia
concentration. Often anaerobic digester sludge dewatering operations occur during the day and
produce return recycle streams high in ammonia concentration (500-1000 mg/L) at times that coincide
with the high influent diurnal ammonia loads. Recycle equalization or treatment helps to provide a more
stable nitrification system and lower effluent NH3-N concentrations.
In many cases, it is advantageous to incorporate a denitrification pre-anoxic step with
nitrification (MLE process) due to the many benefits and improved operational stability. The advantages
include 1) less aeration energy as the nitrate produced can be used for BOD removal, 2) the production
of alkalinity to offset the alkalinity used by nitrification, which in some cases eliminates the need to
purchase alkalinity, and 3) a more stable, better settling activated sludge process as the anoxic-aerobic
processes favor good settling floe-forming bacteria over filamentous growth.
The effluent nitrogen goals greatly affect the process design choices and system operation. For
an effluent goal of 10 mg/L TN, an MLE process is often sufficient for activated sludge treatment with
secondary clarifiers or membrane separation. However, with water conservation leading to more
concentrated wastewater, these processes alone may not be sufficient due to the fact that they are
limited to 80-85% removal of the influent TN.
For TN effluent goals of 3 to 5 mg/L or lower, some form of post anoxic treatment is generally
needed. One option is to convert an MLE process to a Bardenpho process by adding another anoxic-
aerobic set of tanks. Although the endogenous respiration rate of the bacteria can be used to consume
nitrate in the post anoxic tanks, it is often necessary to add an exogenous carbon source. Other
alternatives to using exogenous carbon sources include denitrification filters instead of adding more
activated sludge tank volume, step feed with carbon addition in the last pass, and IFAS processes.
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Denitrification processes require sufficient carbon to drive the nitrate/nitrite reduction
reactions. Characterization of the influent wastewater with regard to its organic strength and soluble
fraction and the TN and ammonia concentrations is needed to fully understand a system's carbon needs.
In addition, design and operating methods that eliminate or minimize DO feeding to anoxic zones can
reduce the amount of exogenous carbon needed and provide a more stable operation. Low DO zones
prior to downstream anoxic tanks or for withdrawal of recycle to preanoxic zones should be considered.
5.5 Impacts on Sludge Production and Handling
It has been documented by both research and full scale experiments that BOD removal by
activated sludge using nitrate as the electron acceptor instead of DO will result in a 20% or more
reduction in waste activated sludge (WAS) production for the same operating conditions. Full-scale
investigations near Melbourne, Australia achieved as high as a 40% reduction in WAS, and
implementation of nitrogen removal at the York River, VA, plant resulted in a reduction of more than
50% in WAS production. The impact this will have on total sludge production by a treatment plant will
depend upon how much waste sludge is produced by other treatment units such as primary clarifiers
and chemical treatment with precipitating chemicals. Additionally, implementation of nitrogen removal
at conventional activated sludge plants can improve the thickening characteristics due to decreasing the
amounts of filamentous bacteria in the activated sludge. If an external carbon source is added to
improve the rate of denitrification, there will be an increase in WAS production compared to when no
external carbon source is added. If an external carbon source is used to supplement denitrification, it is
likely that the small increase in solids production will be offset by endogenous respiration due to longer
SRTs. Solids produced from nitrogen removal processes generally thicken and dewater well and show
no negative impact on any solids processing system.
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6. Biological Phosphorus Removal and Combination Processes
This chapter provides an overview of the principles behind biological phosphorus removal (BPR).
It describes existing configurations that can achieve phosphorus removal and in many cases,
simultaneous nitrogen removal. Key operational issues, impacts on sludge handling, and a summary of
ongoing research related to BPR removal are also provided.
6.1 Principles
Biological phosphorus removal is achieved by contacting phosphorus accumulating organisms
(PAOs) in the RAS with feed, containing volatile fatty acids (VFA), in a zone free of nitrates and DO
(anaerobic zone). Phosphorus is released in this zone providing energy for uptake of VFAs that are
polymerized and stored inside the PAO cells. The anaerobic zone is followed by an aeration zone where
the polymerized VFAs are metabolized and phosphorus is taken up again to store excess energy from
the metabolism. The phosphorus content of the mixed liquor suspended solids (MLSS) would be similar
to that of the waste activated sludge (WAS). When nitrification occurs in the aeration basin, nitrates will
be present in the RAS, resulting in some metabolism of the VFA before storage, thereby reducing
subsequent phosphorus uptake. Some form of denitrification (anoxic zones) must be used to
reduce/remove the nitrates from the RAS. The process flow sheets now known as Pho-redox (A/O) and
3 Stage Pho-redox (A2/O) as well as the modified Bardenpho process were first published by Barnard
(1975) as the Pho-redox flow sheets for the removal of phosphorus. The theory for the functioning of
the PAO was first suggested by Fuhs & Chen (1975) and can briefly be described in conjunction with
Figure 6.1 below.
Fuhs & Chen Theory
PAOs have the ability to store a large mass of phosphorus in their cells in the form of
polyphosphates. Polyphosphates are formed by a series of high-energy bonds. The organisms can
subsequently get energy from breaking these bonds. The polyphosphate globules within the cells
function just like energy storage batteries. The storage of polyphosphates (energy), takes place in the
aeration zone. In the anaerobic zone, these obligate aerobic bacteria can take up short chain VFA such
as acetate and propionate and store them in the form of intermediate products such as poly-B-hydroxy-
butyrate (PHB). The energy for transferring the food across the cell membranes in the anaerobic zone is
derived from breaking phosphorus bonds. Excess phosphates are released to the liquid in the anaerobic
zone. Some magnesium and potassium ions are co-transported across the cell walls with phosphates.
PAOs can only get energy from the food they have taken up in the anaerobic zone when they pass to the
aerobic zone where oxygen is available. They use oxygen to metabolize the stored products, deriving
enough energy to take up all the released phosphates as well as those in the influent, and store them in
the cells. The WAS will contain sufficient phosphate-enriched PAOs to remove most of the phosphorus
from the waste steam once enhanced BPR is established.
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Aerobic
Anaerobic
Figure 6-1 Theory of biological phosphorus removal in activated sludge
The right carbon source, in this case a combination of acetates and propionates, is essential for
BPR. The wastewater characteristics are thus important. In general, it can be said that you need at least
40:1 COD:TP or about 18:1 BOD5:TP in the process influent wastewater to reduce effluent phosphorus to
less than 1.0 mg/L In addition, some of the COD should consist of short chain VFAs. More COD may be
required if nitrates must also be denitrified.
Biological phosphorus removal can work in with or without nitrification. When nitrification
occurs, denitrification within the process is important to reduce the nitrates that may be returned with
the RAS. While the anaerobic zone serves mostly as a contact zone for VFAs and PAOs, some
fermentation of easily biodegradable carbon compounds (rbCOD) to acetate and propionate may take
place. In most plants the readily biodegradable material is in short supply and must be reserved for the
PAOs.
When nitrate or oxygen is discharged to the anaerobic zone, two things may happen, both
undesirable:
• They will prevent fermentation of rbCOD to acetic and propionic acid
• Nitrates or DO could serve as electron acceptors for PAOs and other organisms that will
metabolize the VFA and so deprive the PAOs of the substance that they need to store for
growth and phosphorus removal.
In the absence of electron acceptors such as DO and nitrates in the anaerobic zone, PAOs are
favored to grow since they can take up and store the VFA under anaerobic conditions, thereby making it
unavailable for other aerobic and facultative heterotrophs in the aerobic zone.
Biological removal of both nitrogen and phosphorus at the same WWTP is common. Both
functions require a carbon source. While denitrification organisms can feed on quite a number of easily
degradable materials such as methanol, sugar, glucose, acetate and propionate, PAOs are restricted to
the latter two for polymerization and storage (e.g. adding methanol to the anaerobic zone will reduce
nitrates but not assist in the removal of phosphorus). See section 6.4.1 for detailed discussion on
carbon sources for biological phosphorus removal.
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6.2 Current Configurations
The basic design of anaerobic, anoxic, and aerobic zones, in that order, has been achieved in
many different configurations. The configurations vary in the number of stages, the nature and location
of recycles, and the operation of the process. Each process was modified from the standard biological
activated sludge design in order to accomplish various design goals (e.g., protection of the anaerobic
zone from excess nitrate recycle). The primary processes are listed below. Of these, all will also
biologically remove nitrogen except for the Pho-redox process.
• Pho-redox (A/O)
• 3 Stage Pho-redox (A2/O)
• Modified Bardenpho
• University of Capetown (UCT) and Modified UCT (MUCT)
• Johannesburg (JHB), Modified Johannesburg, and Westbank
• Orange Water and Sewer Authority (OWASA)
• Oxidation ditches with anaerobic zones or phases added
• SBR operated with an anaerobic phase
• Hybrid chemical/biological processes
The performance of these technologies depends on many site specific factors, including but not
limited to temperature, hydraulic and organic loading, recycle rates, and return streams. The
technologies described in this section are generally capable of phosphorus removal to effluent levels
between 0.5 and 1.0 mg/L Operating strategies that can be used to enhance biological treatment and
achieve these and, in some cases, even lower effluent levels are described in Section 6.4.
Biological phosphorus removal can be combined with other technologies to achieve very low
effluent concentrations (< 0.2 mg/L). Chemical addition combined with biological removal of
phosphorus has been used to consistently achieve low levels. WEF and ASCE (1998) recommend that
WWTPs have chemical addition capabilities even for well operating BPR plants to provide backup
phosphorus removal in the event of power outages, pipe breaks, or other unforeseen events.
Solids removal can also be a limiting factor in achieving phosphorus removal below 0.2 mg/L.
Very low phosphorus levels generally require a TSS level of less than 5 mg/L. Tertiary filtration (see
Chapter 7), membrane bioreactors (see Chapter 5), and advanced clarification processes (see Chapter 4)
can achieve TSS levels less than 5 mg/L.
6.2.1 Pho-redox (A/O) and 3 Stage Pho-redox (A2/O)
The Pho-redox (A/O) process is a conventional activated sludge system with an anaerobic zone
at the head of the aeration basin. The RAS is pumped from the clarifier to the anaerobic zone. It is a
low SRT process, operated to avoid nitrification. With no nitrates in the RAS the process is reliable and
easy to operate except at temperatures in excess of 25°C when nitrification is difficult to avoid. The 3
Stage Pho-redox (A2/O) process adds an anoxic zone after the anaerobic zone to achieve de-nitrification.
In addition, a nitrate rich liquor is recycled from the end of the aerobic zone to the head of the anoxic
zone to enhance de-nitrification. A shortcoming of the 3 Stage Pho-redox process is that there will be
nitrates present in the RAS, potentially making the process unreliable.
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Influent
RAS
Anaerobic
Aerobic
Figure 6-2 Pho-redox Process (A/O)
Influent
Mixed Liquor Recycle
Effluent
Anaerobic
RAS
\ Anoxic
Aerobic
Figure 6-3 3 Stage Pho-redox Process (A2/O)
There are existing Pho-redox plants in Pontiac, Ml and Titusville, FL Performance data for these
plants exists in WEF (1998). Performance data on 3 Stage Pho-redox plants in Largo, FL; Fayetteville, AR;
Montgomery County, PA; Warminster, PA; Newark, OH; Goldsboro, NC; Frederick, MD; and Sod Run, MD
are available in WEF (1998) and WEF and ASCE (2006). Neethling et al. (2005) reported data for 3 Stage
Pho-redox plants in Durham, NC and Nine Springs, Wl and for Pho-redox plants in Missoula, MT and
Grand Prairie, Alberta.
6.2.2 Modified Bardenpho
The Bardenpho process removes nitrogen to low concentrations. The addition of an anaerobic
zone at the head of the process enables phosphorus removal as well. The process consists of 5 stages:
an anaerobic stage followed by alternating anoxic and aerobic stages. A nitrate-rich liquor is recycled
from the first aerobic stage, designed for complete nitrification, to the first anoxic stage. The RAS is
recycled from the clarifier to the beginning of the anaerobic zone. Since the nitrates in the RAS ranges
from 1 to 3 mg/L, it does not seriously interfere with the mechanism for phosphorus removal as can
happen in the 3 Stage Pho-redox process.
Performance data on existing plants in Palmetto, FL; Kelowna, BC; Orange County, FL; Fort
Myers, FL; City of Cocoa, FL; Tarpon Springs, FL; Johannesburg, South Africa (Goudkoppies); and
Medford Lakes, NJ are published in WEF (1998) and WEF and ASCE (2006). Neethling et al. (2005)
reported performance data on modified Bardenpho plants at Reedy Creek (Disney World) and Iron
Bridge, FL.
6.2.3 University of Cape Town (UCT) and Modified UCT (MUCT)
The UCT process was designed to reduce nitrates to the anaerobic zone when high removal of
nitrates in the effluent is not required. It consists of three stages: an anaerobic stage, an anoxic stage,
and an aerobic stage. The RAS is returned from the clarifier to the anoxic zone instead of the anaerobic
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zone to allow for denitrification and to avoid interference from nitrate with the activation of the PAOs in
the anaerobic stage. A nitrate rich stream is recycled from the aerobic zone to the anoxic zone.
Denitrified mixed liquor is recycled from the anoxic zone to the anaerobic zone.
UCT Process
LL
Modified UCT Process
Figure 6-4 UCT and Modified UCT Process
Several modifications of the process exist. Sometimes it can be difficult to achieve the level of
denitrification in the anoxic zone required to protect the anaerobic zone from nitrates when the zone is
receiving both RAS and high internal nitrate recycle flows. This problem led to the development of the
modified UCT process, which splits the anoxic zone into two stages. The nitrate rich recycle from the
aerobic zone is recycled to the head of the second anoxic stage. The nitrate containing RAS is recycled
to the first anoxic stage where it is denitrified. Next, the denitrified RAS is recycled from the end of the
first anoxic stage back to the head of the anaerobic stage and mixed with the incoming wastewater. The
Virginia Initiative Plant (VIP) is similar to the UCT process, but the anaerobic and anoxic zones are baffled
into two or more sections each to increase rates of reaction in the first section of each zone, thereby
firmly establishing the desired anaerobic and anoxic condition in the second section.
Performance data for existing VIP plants operated by the Hampton Roads Sanitation District
(HRSD) at Norfolk and Nansemond, VA, for UCT plants in Traverse City, Ml, and Lethbridge, Alberta and
for modified UCT plants in King County South, WA and Kalispell, MT are available in WEF and ASCE
(2006). Performance data are also available for the HRSD VIP plants along with the modified UCT plants
at Kalispell, MT and McDowell Creek, NC in Neethling et al. (2005).
6.2.4 Johannesburg (JHB), Modified Johannesburg and Westbank
The JHB process is similar to the 3 Stage Pho-redox process, but has a pre-anoxic tank ahead of
the anaerobic zone to protect the zone from nitrates when low effluent nitrates are not required. The
low COD of the wastewater limited the de-nitrification capacity in the original plant (Nothern Works),
resulting in nitrates in the RAS. This reduced BPR so much that a pre-anoxic tank was included on the
RAS line to remove the nitrates from the RAS flow using endogenous respiration, before the flow
entered the anaerobic zone. The modified JHB process adds a recycle from the end of the anaerobic
zone to the head of the pre-anoxic zone to provide residual, readily biodegradable compounds for
denitrification.
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The Westbank process is similar to the JHB process but adds some primary effluent to the
anaerobic zone to assist in denitrification with the remainder of the primary effluent being discharged to
the anaerobic zone. During storm flows, excess flow is passed directly to the main anoxic zone. VFA
obtained from acid fermentation of the primary sludge is passed to the anaerobic zone.
Influent
Influent
Johannesburg (JHB) Process
Modified JHB Process
Anoxic LJ Anaerobic ^J Aerobic
Figure 6-5 JHB and Modified JHB Process
Effluent
Westbank Process
Figures 6-6 Westbank Process
6.2.5 Orange Water and Sewer Authority (OWASA)
The OWASA process was developed by adding activated sludge from a biological nitrogen
removal process to a trickling filter plant. Then, nitrified effluent from the trickling filter is fed to the
aerobic zone of the activated sludge system. Because the VFAs have been destroyed by the trickling
filter, it is necessary to ferment the settled organic solids from the primary clarifier to produce sufficient
VFAs for BPR. Next, the fermented supernatant is passed to an anaerobic (nutrition) zone and mixed
with the RAS to initiate BPR.
Mixed liquor then flows from the nutrition zone to an anoxic zone and then to an aerobic
zone. Alternatively, simultaneous nitrification and denitrification takes place in the aeration zone, as
shown in Figure 6-7. Performance data for the original plant in Orange County, NC, are available in WEF
and ASCE (2006).
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|::| Nutrition Zone ^ Anoxic |~|Aeroblc
Figure 6-7 OWASA Process
6.2.6 Oxidation Ditches
There are several oxidation ditch designs that can remove phosphorus. They normally consist
of an anaerobic zone ahead of the oxidation ditch whereas simultaneous nitrification and denitrification
takes place within the ditches. Oxidation ditches typically operate as racetrack configurations around a
central barrier, with forward mixed liquor flows of approximately 1 foot per second or more. It is
possible, by manipulating the DO transferred to the mixed liquor, to establish both anoxic, aerobic and
near anaerobic zones within the racetrack configuration, even though the high flow velocities
accomplish complete mixing of the wastewater with the RAS.
There are many forms of oxidation ditches, such as the Carousel, the Pasveer Ditch and the
Orbal process. The Orbal process creates anaerobic and anoxic zones in the outer of three concentric
oval shaped ditches with the RAS recycled from the clarifier to the anoxic zone. It is also possible to
introduce an anaerobic tank before the ditch to accomplish BPR in the combined system. The Pasveer
Ditch and the Carousel system also can be used in conjunction with an anaerobic zone to accomplish
BPR, in addition to simultaneous nitrification and denitrification within the ditches. Because of the very
high internal recycle within the ditches, very low nitrate concentrations can be achieved in the mixed
liquor before settling, and anaerobic conditions are easy to maintain in the anaerobic zone, thereby
resulting in efficient BPR. The layout would resemble a Pho-redox process with simultaneous
nitrification-denitrification (SND) in the aeration basin. Alternatively the Carousel or Pasveer Ditch could
be used as the aeration stage in either the 3 Stage Pho-redox or the Modified Bardenpho process.
The VT2 process at Bowie, MD, operates two Pasveer ditches in series with dedicated anoxic,
near anaerobic and aerobic zones. It also has a side stream anaerobic zone that receives only 30 percent
of the influent flow to enhance BPR. Denitrified MLSS for the anaerobic zone are obtained from the end
of the near anaerobic zone of the adjacent ditch. Operated without primary sedimentation, the system
consistently obtains very low (<0.25 mg/L) effluent TP without chemicals or effluent filtration. The
ditches are operated in series because the plant has limited clarification capacity, and series operation
results in lower MLSS concentrations to the clarifiers. The biodenipho process also uses pairs of ditches.
The ditches in the biodenipho process operate in alternating anoxic-aerobic modes. An anaerobic tank is
placed before the ditches for BPR and the ditches are alternated between nitrification and de-
nitrification.
Performance data for existing plants at Bowie, MD and North Gary, NC are available in WEF and
ASCE (2006).
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6.2.7 Sequencing Batch Reactors (SBR)
SBRs are fill-and-draw reactors that operate sequentially through the various phases by means
of adjusting the mixing and aeration. The reactor phases can be set and automated to allow the mixed
liquor to go through an anaerobic/anoxic/aerobic progression as is necessary for removal of phosphorus
and nitrates. Because of the fill-and-draw nature of SBRs, it actually is necessary to remove the nitrates
remaining from the previous cycle before anaerobic conditions can be established, thus the typical
treatment progression becomes anoxic/anaerobic/aerobic. However, SBRs are almost always operated
without primary sedimentation, so they still usually have a favorable BOD5:TP ratio for effluent TP of
somewhat less than 1.0 mg/L during the settling phase.
Performance data for plants in Oak Point, Ml; Grundy Center IA; Culver City, IN; Armada, Ml; and
Manchester, Ml are in WEF and ASCE (2006).
6.2.8 Hybrid Chemical / Biological Processes
The PhoStrip configuration, used mainly in non-nitrifying plants, pulls a side stream off the RAS
in a conventional activated sludge plant. The side stream is concentrated and retained for a day or more
in a thickening tank where the solids blanket is deep enough to produce anaerobic conditions and
fermentation, resulting in the release of phosphates by the microorganisms. Lime is then added to the
supernatant stream to precipitate and remove phosphate. The thickened, fermented sludge is passed
back to the main aeration basin. Existing plants include Seneca Falls, NY; Lansdale, PA; Adrian, Ml;
Savage, MD; Southtowns, NY; Amherst, NY; and Reno-Sparks, NV.
The Biological Chemical Phosphorus and Nitrogen Removal (BCFS) configuration is similar to the
modified UCT process. In this process, a sludge stream is removed from the anaerobic zone. Ferric
chloride is added to the sludge thickener to remove phosphate. This provides an advantage over
chemical addition to the secondary clarifier because it does not require the chemical sludge to be
recycled. There is an existing plant at Molten in the Netherlands (WEF and ASCE, 2006), but no
performance data are available.
6.3 Emerging Technologies
Many plants that are not specifically configured for BPR nevertheless achieve phosphorus
removal to less than 1 mg/L. The first such observation in a nitrifying plant was in a four-stage
Bardenpho plant where mixed liquor was recycled from the second anoxic zone to an unstirred
fermenter, then returned to the anoxic zone. The CATABOL™ and Cannibal Processes claim to reduce
excess secondary sludge production by passing mixed liquor or RAS through an anaerobic (fermenting)
stage and then back to the main stream aeration system. In addition, both processes pass the mixed
liquor through a process for removal of some of the inert solids. Both processes claim to get similar
phosphorus removal to that for the Bardenpho plant described above. All of these processes rely on the
fermentation of some of the mixed liquor for producing VFA that assists in the biological removal of
phosphorus. The Town of Gary, NC, has been using a system by which some of the sludge in the return
streams of a biological nitrogen removal plant is subjected to anaerobic conditions similar to that of the
other processes described above resulting in an effluent phosphorus concentration of less than 0.5
mg/L.
There is a similarity between these processes and ad hoc processes for switching off aeration in
plug-flow plants for promoting phosphorus removal. These ad hoc processes take various forms. The
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Piney Water, CO, plant is a 5-stage Bardenpho plant with no primary sedimentation and little VFA in the
influent, which resulted in little phosphorus removal. By switching off a mixer in one of the anaerobic
zones, sludge settled to the bottom and fermented, which supplied the VFAs for reducing the ortho-
phosphorus to less than 0.2 mg/L A similar operation at the Henderson, NV, plant in a JHB type process
had the same effect. Some plug-flow aeration plants succeeded in reducing phosphorus to below 1
mg/L by turning off aeration at the feed end of the plant, such as the Blue Lakes and Seneca plants
operated by the Metropolitan Council Environmental Service in Minnesota and the St. Cloud, MN, plant.
The Joppatowne plant operated by Harford County, MD, consists of an MLE plant with some sludge
accumulation in the anoxic zone while reducing the phosphorus from 7 mg/L in the influent to around 1
mg/L in the effluent.
All of these plants use the same principle of fermenting some of the mixed liquor sludge or
underflow from the final clarifiers, either inside the main stream tanks or in a side stream basin. There
are many instances where enterprising operators can achieve 80 percent or more phosphorus removal
by turning off air or mixers in conventional treatment plants.
There is a Catabol plant in Cartersville, GA (USEPA, 2008a); however, there are no published
data for this plant.
6.4 Operational and Design Considerations
Important factors that affect BPR include:
• Bioavailable COD:P ratio in the anaerobic zone influent, including adjustments by VFA
addition and sludge fermentation
• SRTandHRT
• Presence of oxygen or nitrate in the anaerobic zone
• Backmixing of oxygen
• Temperature
• pH
• Secondary release under anaerobic conditions
• Sufficient oxygen in the aerobic zone
• Inhibition
• Flow and load balancing
6.4.1 COD:P Ratio
The PAOs need VFAs in the form of acetic and propionic acid. These acids may be in the feed or
can be produced through fermentation of soluble rbCOD such as sugar, ethanol, etc., in the anaerobic
zone. As a rough estimate of the propensity for phosphorus removal to an effluent concentration less
than 1.0 mg/L, the COD:P ratio typically should be about 40 or more.
VFA is produced through fermentation of municipal wastewater or it can be added as a
commercial or waste product. Some wastewater collection systems that are relatively flat and have long
collection times may have sufficient fermentation in the collection system to provide the necessary
VFAs, but it will vary monthly depending upon the temperature and flow conditions in the collection
system. Force mains are excellent fermenters for the production of VFA. Systems that do not have a
COD/P ratio of at least 40 will most likely need to supplement VFAs to achieve effluent phosphorus
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concentrations below 1.0 mg/L. However, they will still achieve substantial BPR with lower ratios if
appropriately operated. See below for a more detailed discussion of VFAs.
Recent studies suggest that the instantaneous COD:P ratio is more important than the overall
average (Neethling et al., 2005). Short term drops in the BOD:P ratio in the primary effluent to below
that required for sustainable phosphorus removal correlated well with rises in effluent phosphorus.
Intermittent recycles of phosphorus rich return streams may cause short term variability in the BOD:P
ratio. Controlling or eliminating these recycles can improve plant performance. Weekend changes in
the BOD:P ratio also can affect performance.
Another group of organisms, glycogen accumulating organisms (GAOs), also has the ability to
take up acetate in the anaerobic zone, not by using energy in phosphate bonds but by using stored
glycogen as the energy source. Under certain conditions, such as high temperatures or low phosphorus
concentrations relative to the influent bioavailable COD, they may out-compete PAOs for the VFAs,
which would result in less or no release of phosphorus in the anaerobic zone. This in turn will result in
less or no overall phosphorus removal. GAOs use the stored energy in the form of glycogen to take up
VFAs and store them as a complex carbohydrate containing poly-hydroxy valerate (PHV), instead of PHB
formed with poly-phosphorus as the energy source. When this begins to happen, there is a slow decline
of phosphorus removal by the biological system.
There is still a debate amongst researchers about the conditions likely to favor GAOs over PAOs.
Summarizing a number of publications, it would appear that the following conditions favor the growth
of GAOs over that of PAOs:
• High SRT
• High temperature over 28 °C
• Longer non-aerated zones
• Stronger wastes with low TKN content
• Periods of intermittent low BOD loads
• If the VFA consists mostly of either acetate or propionate
• Polysaccharides such as glucose are fed to the anaerobic zone.
• Low pH in the aerobic zone
Further confirmation is needed for some of these factors.
Volatile Fatty Acid Addition
Only VFAs such as acetic and propionic are taken up by PAOs. Reported doses of VFA for
successful phosphorus removal range from 3 to 20 mg/L VFA per gram of phosphorus removed. These
numbers, however, do not take into account the rbCOD that is fermented in the anaerobic zone. It is
more accurate to look at the rbCOD/P ratio for good phosphorus removal, which ranges from 10 to 16.
(Barnard, 2006). Surveys show that it is rare for a WWTP treating municipal sewage to achieve more
than 95 percent removal of phosphorus by biological processes without adding VFAs (Neethling et al.,
2005).
An Australian study shows that while both PAOs and GAOs could use acetate, PAOs will have a
competitive advantage when the VFAs consist of roughly equal parts of acetic and propionic acid as a
growth medium. PAOs that are fed on acetate are able to switch to propionate much more quickly and
effectively than GAOs (Oehmen et al., 2005). This finding led to a strategy to feed equal amounts of
acetic acid and propionic acid as the optimal for stimulating PAO growth (Oehmen et al., 2006, Bott et
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al., 2007). One study shows that isovaleric acid drives BPR even better than acetic acid (Bott et al.,
2007). Isovaleric acid, however, is much more expensive than acetic acid and is more odorous. It also is
not significantly generated in the primary sludge fermentation process. Addition of rbCOD such as
sugars and alcohols containing two carbons or more can increase phosphorus uptake by PAOs when
added to the anaerobic zone but may cause sludge bulking if dosed in excess (Jenkins and Harper, 2003).
Sludge Fermentation
Anaerobic fermentation produces VFA consisting mainly of acetic and propionic acid. Some
configurations, such as the Westbank and OWASA configurations, make use of anaerobic fermentation
of the primary sludge to provide VFAs to the nutrient removal process. A fermentation process,
however, can be added to any configuration to provide VFAs, especially in areas where little
fermentation takes place in the collection system. Fermentation of the primary sludge or the RAS will
produce VFA. Primary sludge fermentation is used more frequently.
There are several primary sludge fermenter designs that can accomplish this. The simplest
configuration involves allowing the formation of a thicker sludge blanket in the primary clarifier itself
and returning some of the thickened sludge to either the primary clarifier or to a mixing tank ahead of
the primary clarifier to allow elutriation of the VFA to the primary effluent. This is referred to as an
activated primary sedimentation tank (Barnard, 1984). Another variation is to pump some sludge to a
complete-mix tank ahead of the primary clarifier, to accomplish fermentation. The sludge is then passed
to the primary clarifier for elutriation of the VFA. Both of these processes lead to an increased load on
the primary clarifier and some VFA may be lost due to aeration between the primary clarifier and the
anaerobic zone. Sludge age should also be controlled to prevent methanogenic bacteria from growing
and converting the VFA to methane. Usually, a SRT less than 4 days is sufficient for this.
Alternative methods accomplish fermentation in a gravity sludge thickener by holding the sludge
under anaerobic conditions for 4 to 8 days. The supernatant can then be fed directly to the anaerobic
zone and a high load on the primary clarifier can be avoided. Thickening can either be accomplished
with a single thickener or in two stages. The two-stage process can either be a complete mix tank,
followed by a thickener or two thickeners in series. It has been shown that adding molasses or other
sources of readily biodegrable COD can improve the performance of fermenters (Bott et al., 2007).
RAS can also be fermented in a side stream process. The fermentation zone is similar to the
anaerobic or anoxic zone of many biological processes. RAS fermentation could be used in any BPR
process, but is most common in processes without primary clarifiers.
Research and experience have revealed some key design considerations for primary fermenters
(WEF and ASCE, 2006). These processes can have high solids content and may need a positive
displacement pump to operate properly. Because fermentation can lower the pH and produce carbon
dioxide and hydrogen sulfide, corrosion resistant materials should be used. Odor control may also be
necessary if hydrogen sulfide is produced. Monitoring of pH and oxidation reduction potential (ORP)
may be desirable to control the process.
6.4.2 Retention Time
The concentration of phosphorus in the sludge typically increases as the SRT increases, although
the impact is very small over the SRT range of 4 to 30 days. Efficient phosphorus uptake typically
requires a minimum SRT of 3 to 4 days depending on temperature. Higher SRTs will not increase
phosphorus uptake, given there is sufficient VFAs available. If SRT becomes too great, however, effluent
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quality can degrade. This can be due to release of phosphorus as biomass degrades (WEF and ASCE,
2006).
Both anaerobic and aerobic HRT can affect the amount of phosphorus stored by PAOs.
Sufficient time should be allowed for the formation of VFAs and storage of the polyhydroxyalkanoates
(PHAs) in the anaerobic zone, although the reactions are relatively fast. If the time is too short,
phosphorus uptake in the aerobic zone will be lower than achievable because insufficient PHAs were
stored in the anaerobic zone. It has been reported that the ratio of HRT in the anaerobic zone to the
HRT in the aerobic zone is important. One study found that a ratio of between 3 and 4 for aerobic HRT
to anaerobic HRT led to optimal plant operation (Neethling et al., 2005).
6.4.3 Temperature
High temperatures can have an adverse effect on phosphorus removal. At temperatures greater
than 28°C, phosphorus removal will generally be impaired, apparently by the predominance of the GAOs
(Bott et al., 2007). At the low end of the temperature scale, Erdal et al. (2002) found that PAOs
outcompeted GAOs at 5 °C even though the PAO metabolism was slower at 5 °C than at 20 °C. The GAOs
virtually disappeared in the 5 °C reactor.
Modeling studies have shown that GAOs can predominate at higher temperatures because of
their increased ability to uptake acetate at those temperatures compared to PAOs (Whang et al., 2007).
Low temperatures can also lower phosphorus uptake but have been shown to not be an issue in well
operated and properly acclimatized plants (WEF and ASCE, 2006).
6.4.4 Presence of Oxygen or Nitrate in the Aerobic Zone
If oxygen or nitrate is present in the anaerobic zone, organisms that use oxygen or nitrates as
electron acceptors will preferentially grow by fully oxidizing the organics to CO2 and H2O, thereby
reducing the VFAs available for polymerization and storage by the PAOs. Nitrate can also inhibit
fermentation of rbCOD because most of the fermenters are facultative and can use the nitrate as an
electron acceptor to fully oxidize the rbCOD instead of producing VFAs as an end product of
fermentation, thus depriving the PAOs of organics they can polymerize and store. Therefore, recycle of
streams containing high DO and nitrate concentrations to the anaerobic zone should be avoided.
Introduction of oxygen through pumps and other devices should also be avoided.
6.4.5 Avoiding Backmixing of Oxygen
Another potential source of oxygen and nitrates to the anaerobic zone is backmixing from
downstream zones. In configurations where the anaerobic zone is followed immediately by an anoxic or
aerobic zone, backmixing can cause elevated concentrations of nitrates and/or DO in the anaerobic zone
leading to favoring of organisms other than PAOs. The problem can be avoided by increased baffling or
changing the mixing rates. This problem is more likely to occur when the downstream zone is aerated,
because aeration of mixed liquor increases the liquid depth, making the liquid level in the aerobic zone
higher than in the non-aerated zone.
6.4.6 pH
Low pH can reduce and even prevent BPR. Below pH 6.9 the process has been shown to
decline in efficiency (WEF and ASCE, 2006). This is possibly due to competition with GAOs. Filipe, et al.
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(2001), found that GAOs grow faster than PAOs at a pH of less than 7.25. Because many wastewater
processes such as chemical addition and nitrification can lower pH, this should be monitored and
adjusted if necessary. It also has been shown that it is not possible to establish enhanced biological
phosphorus removal (EBPR) when the pH is less than 5.5, even though an abundant amount of acetic
acid is present in the anaerobic zone (Tracy and Flammino, 1987; Randall and Chapin, 1997).
6.4.7 Anaerobic Release
Secondary release of phosphorus occurs when the PAOs are under anaerobic conditions in the
absence of a source of VFA. The energy stored as polyphosphate is used for cell maintenance and
phosphorus is released to the liquid phase (Barnard, 1984). There will then be no stored food to supply
energy for the uptake of phosphorus upon subsequent aeration. This may occur in the following process
stages:
• In the anaerobic zone if the retention time is too high and the VFA is depleted well within
the required retention time.
• In the main anoxic zone when that runs out of nitrates.
• In the second anoxic zone as shown in Figure 6.8 when there are no nitrates to be removed.
• In the sludge blankets of final clarifiers when the RAS rate is too low and sludge is not
removed fast enough.
Additionally, release may happen in aerobic zones that are too large, resulting in stored
substrate depletion and destruction of PAO cells by endogenous metabolism.
Since there was no food storage associated with the phosphorus release, additional carbon is
then required to take up the phosphorus released, but the amount in the influent may be insufficient.
Therefore, chemicals must be added to remove the excess phosphorus. Over-design of biological
nutrient removal systems could thus lead to a higher demand for an external source of VFA.
50
"I
o Anaerobic
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20
Figure 6-8 Example of Secondary release in Second Anoxic Zone
Phosphorus will be released in sludge treatment processes that are anaerobic. Gravity
thickening of BPR sludge can lead to phosphorus release if long retention times are used. Using
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mechanical dewatering instead of gravity dewatering allows less retention time and less phosphorus
release (Bott et al., 2007). It is usually recommended that dissolved air flotation (DAF) be used to
thicken BPR sludge to reduce the amount of phosphorus release. DAF thickening can be quite successful
for the reduction of release, but if the thickened sludge is left on the DAF beach too long before
removal, excessive release will occur, just as it will when the sludge is left too long in a gravity thickener.
Anaerobic digestion will also lead to phosphorus release although some phosphorus will be
precipitated as either a metal salt (e.g. calcium phosphate) or as struvite (magnesium ammonium
phosphate, MgNH4PO4). BPR sludge takes up and releases magnesium along with phosphates, and these
two ions combine with ammonium, also present in abundance in anaerobic digesters, to form struvite.
Struvite formation is very fast, and will continue until one of the three ions is reduced to that ion's
solubility level. Magnesium is usually present in the lowest concentration, and its depletion typically
limits struvite formation within the anaerobic digester. Calcium phosphate precipitates also tend to
form in anaerobic digesters, but they form much more slowly than struvite and the formation tends to
be non-stoichiometric. If substantial amounts of phosphates are precipitated by calcium along with the
struvite formation, there will be little if any propensity for struvite to form when the sludge exits the
anaerobic digesters. Also, if the digested sludge is composted after dewatering, the resulting Class A
sludge will be enriched in magnesium, phosphorus, nitrogen, and, to a lesser extent, potassium, which
also is taken up and released with phosphorus by PAOs. Thirty percent of the phosphorus entering the
anaerobic digesters at the York River plant during BPR experimentation was recycled back to the
headworks from belt filter press dewatering (Randall et al., 1992).
Alternatives to anaerobic digestion such as composting, drying, or alkaline treatment can be
used to reduce phosphorus release. There have been several studies which have examined using
struvite precipitation as a way of recovering phosphorus from supernatant from digesters. These
processes have been tested on full scale facilities in Treviso, Italy and Vancouver, Canada (SCOPE, 2004).
When anaerobic release of phosphorus occurs, recycling these streams can overload
phosphorus removal processes. The effect can be worsened when the waste handling process is only
operated intermittently. In some instances there is a high degree of phosphorus precipitation in the
anaerobic digesters and with sufficient VFA in the influent the returned phosphorus may be removed.
However, in most circumstances, some chemicals need to be added to the return streams or to the
anaerobic digester itself so that the metal precipitate will be removed with the dewatered sludge.
6.4.8 Sufficient Oxygen in the Aerobic Zone
It is necessary for oxygen to be present in the aerobic zone for phosphorus to be taken up and
retained in the activated sludge. Maintaining a sufficiently high DO transfer in the aerobic zone
enhances process stability and has been found to be a key factor in phosphorus removal. (Bott et al.,
2007)
6.4.9 Inhibition
EBPR, like any biological process, can be inhibited by chemicals toxic to the organisms. Although
not as sensitive to inhibition as nitrification and rare in practice, the BPR process can be inhibited by
toxic chemicals, including high concentrations of acetate (Randall and Chapin, 1997).
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6.4.10 Flow and Load Balancing
Flows and loads to wastewater treatment plants can vary widely because of regular diurnal use
patterns and because of larger, more irregular disturbances such as storm events. Peaks in either flow,
or nutrient load can stress the system and cause poor performance. Peaks can be evened out using
equalization tanks to balance the flow. Equalization tanks in combination with nutrient sensors can also
be used to balance nutrient loads. In this case, recycle streams high in nutrient concentrations such as
digester supernatant can be stored during peak nutrient loads and recycled during times when
concentrations are lower.
6.5 Impacts on Sludge Handling and Removal
Stored phosphorus adds dry weight to the sludge; however, the actual PAO VSS production will
be less because the reaction is less efficient than heterotrophic metabolism using DO as the electron
acceptor.
Sludge from BPR will be similar to sludge from conventional activated sludge plants, although it
will have a higher phosphorus content. Varying results have been found with some plants reporting
little or no change in settling and dewatering (Knocke et al., 1992) and others reporting enhanced
settling and dewatering properties (Bott et al., 2007). The sludge produced from the process will also
have higher magnesium and potassium concentrations due to co-uptake of these elements with
phosphorus.
Struvite can precipitate in anaerobic processes. With abundant phosphorus and ammonia it is
usually only the magnesium that is in short supply. Some magnesium is released from the digested cells
with the phosphorus and may increase struvite precipitation. Some processes have proposed
precipitating out struvite or other phosphate solids to avoid phosphorus return in recycle streams (Bott
et al., 2007). The struvite crystals, however, depending upon where they form, can plug centrifuge ports,
and pumps and pipes used to convey the sludge, if not controlled. Plugged lines are very difficult to
clean.
6.6 Guidance for Selecting Process Modifications
If an existing activated sludge WWTP needs to lower phosphorus levels in its effluent, a number
of options are available. Some key considerations are summarized below.
For systems that do not have BPR, an anaerobic zone can be added at the head of the plant.
This may be achieved by switching off aerators at the head of the reactor or by adding a separate
reactor. Mixing in the anaerobic zone should be sufficient to retain biological solids in suspension
without introducing oxygen. If baffling is not already present, it could be added to achieve separation of
the anaerobic and aerobic zones. Note that baffling is essential to prevent backmixing because the
liquid level in the aerated zone will always be higher than that in the non-aerated zone. Therefore, an
overflow baffle should be used between zones. Considerations should also be made for additional
pumping needed for any recycle streams. Proper sizing of the anaerobic zone is important to ensure
sufficient VFA is formed and taken up in the aeration basin. If an aerobic zone is converted to an
anaerobic zone, care should be taken to ensure that the remaining aerobic zone is sufficiently sized to
achieve treatment objectives. This usually is not a problem because the anaerobic zone seldom needs
to be more than 15 percent of the total volume, and can be considerably less if fermentation is practiced
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or VFA are added. Note that much of the BOD in typical municipal sewage will be removed from
solution in the anaerobic zone, and this reduces the required size of the aerobic zone, even though most
of the stored BOD will be stabilized in either the anoxic or aerobic zone, or both.
For plants that already have BPR but need additional phosphorus removal, the designers should
start by identifying areas that may be limiting the current process. For example, if recycle streams are
intermittent, overloading of the process may occur during recycle and the process performance may
suffer. Flow equalization to enable constant recycle flows may be an option in these cases. RAS when
returned to the anaerobic zone may introduce nitrates or oxygen that will interfere with PAO
performance. The phosphorus content of the return streams could be reduced by adding some
chemicals to precipitate some of the phosphorus. Reducing oxygen introduction to the anaerobic zone
from upstream processes may be needed to optimize phosphorus removal.
Plants looking to improve phosphorus removal performance should also closely examine the
plant for secondary release of phosphorus. If sludge blankets in clarifiers are too deep, anaerobic
conditions can develop and cause secondary phosphorus release. This can be minimized by using
deeper clarifiers, maintaining low sludge blankets, and increasing the RAS rate, so that the released
phosphorus is pumped from the bottom of the clarifier rather than flowing over the effluent weir.
Sludge handling can also cause excessive phosphorus release such as in gravity thickeners, DAFs and
anaerobic digesters. If supernatant from these processes when poorly managed is recycled, it can
overload the process. Options in this case would be to eliminate the recycle, improve operation of the
process, change the process, or treat the recycle stream to remove phosphorus before it is returned to
the plant.
Another area to examine in seeking improved phosphorus removal is the COD:P ratio. If the
ratio is low, supplementing the current process with VFAs may provide additional removal. VFAs can
either be added as a chemical addition process or produced through fermentation of primary or
secondary sludge.
Other ways of improving TP removal include filtration and chemical addition. Phosphorus is
often attached to colloidal particles and very low phosphorus levels usually require removal of TSS.
Membrane bioreactors (MBR) in combination with biological and/or chemical phosphorus removal can
result in very low effluent levels due to enhanced solids removal. Chemical addition with or without
filtration can also achieve low phosphorus levels.
6.7 Ongoing Research
Research into BPR is ongoing. The Water Environment Research Foundation has committed to
several research goals considered high priority in this area. These include research on selecting the best
external carbon sources and analytical methods for low phosphorus detection. Other research is being
conducted into the competition between PAOs and GAOs and the conditions favoring PAO growth.
Research into carbon augmentation is also ongoing.
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7. Effluent Filtration
Effluent filtration in combination with chemical precipitation can be used to remove
phosphorous down to very low levels (< 0.1 mg/L). USEPA Region 10 (2007) found that 2-stage filtration
through use of a first and second stage filter or by providing tertiary clarification prior to filtration,
resulted in the lowest effluent phosphorus concentrations of 23 WWTPs evaluated. Effluent filtration
can also be used to remove soluble organic nitrogen that is not removed through biological treatment or
settling.
A wide variety of filter types have been used for wastewater treatment, including:
• Conventional down-flow filters
• Deep-bed down-flow filters
• Continuous backwashing upflow sand filters
• Pulsed bed filters
• Traveling bridge filters
• Fuzzy filters
• Discfilter
• Cloth media disk filters
• Membranes
• Blue PRO™ process
• Pressure filters
This chapter describes the various filters listed above, presents key design and operating principles, and
summarizes ongoing research and emerging technologies in this area.
7.1 Types of Filters
7.1.1 Conventional Down-flow Filters
These filters consist of fixed-media beds typically up to 3 feet in depth and are similar to filters
used to treat drinking water. Media can be single media, dual media, or multi media. Single media is
typically sand or anthracite. Dual media combines anthracite and sand. Multi-media filters include a
layer of garnet or ilmenite. Flow in these filters is by gravity from the top down. Most of the removal
occurs in the top few inches of the media. The filter must be taken off-line periodically to backwash the
filter to prevent clogging and too high of a pressure loss.
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Performance data on plants in Breckenridge (Farmers Korner), CO; Dillon, CO; Hillsboro, OR;
Tigard, OR; Alexandria, VA; Occoquan, VA; and Fairfax, VA can be found in USEPA Region 10 (2007).
7.1.2 Deep-bed Down-flow Filters
These filters are similar to conventional down-flow filters but have deeper beds and larger
media size. This gives the advantage of longer run times between backwashes. The size of the media is
limited by the ability to backwash the filter. Because these filters are more difficult to backwash, air
scour is necessary to fully clean the filter bed.
7.1.3 Continuous Backwashing Upflow Sand Filters
During operation of the continuous backwashing upflow filter, water is introduced through
risers at the bottom of a deep sand bed. Water flows upward through the sand bed and over an
overflow weir. Sand and trapped solids flow downward through the filter and are drawn into the
suction of an airlift pipe in the center of the filter. As the sand travels up the airlift pipe, energy from the
air scours the particles and separates the sand from filtered solids. At the top of the airlift pipe, the
clean sand settles back onto the top of the filter and the solids are carried away into a reject line.
These filters have the advantage of having no moving parts other than the air compressor and
requiring less energy and maintenance than traditionally backwashed filters. They are sometimes
referred to by the trade name Dynasand.
EPA Region 10's (2007) report on plants with LOT phosphorus removal includes performance
data for these filters in plants in Aurora, CO; Breckenridge (Iowa Hill), CO; Stamford, NY; and Walton, NY.
7.1.4 Pulsed Bed Filters
Pulsed bed filters are shallow filters with an unstratified fine sand media. An air pulse disturbs
the media and allows penetration of solids into media bed, allowing the entire filter bed to be used for
removal of solids. The pulse is designed to expand the filter operation and reduce the number of
backwash cycles, although the filter must still be periodically backwashed to remove the solids.
7.1.5 Traveling-Bridge Filters
Traveling-bridge filters consist of long shallow beds of granular media. Wastewater is applied to
the top of the media and flows downward. Each cell is individually backwashed by a traveling-bridge
while the other cells continue to operate. The bridge uses filtered water to backwash the filters and
includes surface wash to breakup matted solids or clumps of solids.
EPA Region 10's (2007) report on plants with LOT phosphorus removal includes performance
data for the filters in plants in McMinneville, OR and Milford, MA.
7.1.6 Fuzzy Filters
The fuzzy filter uses a proprietary synthetic filter media that is highly porous. Water flows not
only around the media but also through it, allowing much higher filtration rates. The media is held in
place by a metal plate and flow is from the bottom of the bed upwards. The filter is backwashed by
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raising the plate and introducing a horizontal air stream from alternating sides causing the media to roll
back and forth. The effluent is returned to the plant.
Plants in the following locations have fuzzy filters: Columbus, GA (CSO); Clayton County NE, GA;
Yountville, CA; Rogersville, MO; Golden Poultry/Golden Kist, NC; Orange County, CA; and King County,
Seattle, WA. Performance data for these plants were not available.
7.1.7 Discf liters
Discfilters are a series of parallel mounted disks used to support a cloth filter media. Water
enters a central tube and flows out between the two layers of cloth in each disk. The disks rotate and
are normally 60 to 70 percent submerged. The portion above the water is backwashed using spray
nozzles.
7.1.8 Cloth Media Disk Filters
The cloth media disk filter is similar to the discfilter listed above. In this case the water flows
from the outside of the partially submerged cloth disks and into a center pipe. Disks continue to rotate
during backwash and water is sucked into the disc using suction heads.
7.1.9 Membranes
Membrane systems use a pressure head to drive water through a permeable membrane.
Membrane filters are typically classified by their pore size which in turn determines the size of the
particles they exclude. Microfiltration, ultrafiltration, nanofiltration, and reverse osmosis (RO) remove
increasingly smaller particles. Microfiltration and ultrafiltration remove 3 to 6 logs of bacteria, 95
percent or more BOD, along with most particles (WEF, 2006). Nanofiltration removes nearly all particles
including some viruses. RO removes all particles as well as most large dissolved constituents. The
energy cost for applying the pressure head and the need to replace membranes make membrane
filtration a more expensive technology. It can achieve very low concentrations of nutrients and other
contaminants, however, and is common in water re-use projects.
Membranes can be configured a number of ways including hollow fiber, spiral wound, plate and
frame, cartridge, or in pressure vessels. Membranes can foul from organics, biological activity, or metals
in the wastewater. Typically the water must be pre-treated before using these membranes. Pre-
treatment could be conventional filters, cartridge filters, or larger membrane filters. Disinfection may
also be required to prevent biological fouling.
EPA Region 10's (2007) report on plants with phosphorus removal includes performance data
for the filters in plants in Clifton, CO; Grand Gorge, NY; Ashland, OR; and Parker, CO.
7.1.10 Blue PRO™ Process
The Blue PRO™ process uses a continuous backwashing filter that is designed remove
phosphorus. Filters can be run in series for even greater removal. The filter media (sand) is coated with
a hydrous ferric oxide coating, which enhances phosphorus removal through adsorption. A ferric salt is
added prior to the filter to aid in coagulation and to replace the ferric coating which is abraded from the
sand. Water flows up through the filter while the sand travels down. An airlift tube at the bottom of
the filter carries the sand upward. Turbulence from the compressed air knocks accumulated iron and
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phosphorus along with any solids off the particle as it travels upward. The iron, phosphorus, and
particles are wasted, while the clean sand is deposited on the top of the bed. The filters can be run
biologically active to achieve denitrification.
The Blu-CAT process combines the Blu-Pro process with addition of advanced oxidants. Early
pilot tests show that this process is capable of removing other emerging contaminants along with
phosphorus and microorganisms (USEPA, 2008a).
Performance data are available from a plant in Hayden, OH (USEPA Region 10, 2007) and from a
demonstration plant in Florida (Blue Water Technologies, 2008).
7.1.11 Pressure Filters
Pressure filters are similar to conventional media filters except they are contained in closed
containers and are filtered under pressure. The increased pressure creates a greater head loss and
allows longer times between backwashes.
7.2 Design and Operating Principles
Filtration is mainly affected by the concentration and size distribution of particles entering the
filter. Turbidity is often used as a surrogate for particle concentration. The concentration of particles
will affect run-time in filters and will also affect the required surface area to achieve the desired
filtration. The size distribution of the particles and its relevance to pore size of the granular or
membrane filters will affect the removal mechanisms.
Filtration rate is also an important design parameter. Too fast of a filtration rate can cause floe
to break up and pass through the filter. The optimal filtration rate depends on floe strength, which in
turn depends on the biological treatment processes prior to filtration (e.g., Higher SRTs lead to weaker
floes). The filtration rate, along with the loading rate will determine the area of the filter required. The
higher the loading rate, the more frequent backwashes will be required and the greater the head loss
across the filters. Typical filtration rates are 5 to 15 meters per hour for gravity filters and up to 20
meters per hour for pressure filters (WEF and ASCE, 1998).
Addition of polymers or other coagulant aids can greatly aid filtration. Typical doses for filter
influent are 0.05 to 0.15 mg/L of organic polyelectrolyte (WEF and ASCE, 1998), although jar tests are
conducted to determine the proper dose. Too low a dose can allow uncoagulated particles through the
filter and too high a dose can lead to mudballs and filter clogging.
There are several ways the flow rate can be controlled in filters. Constant-rate fixed head
filtration maintains a constant flow through the filter. This will lead to an increased head above the
filter as the filter run progresses. In constant-rate variable head filtration the rate is kept the same and
the filter is backwashed when the head reaches a certain value. In variable-rate filtration, the rate of
filtration decreases throughout the filter run until it reaches a minimum value and is backwashed.
Variable-rate filtration is less common than constant-rate filtration.
Proper backwashing is also important to filter operation. Without proper backwashing there
can be breakthrough of particles and turbidity. Lack of a proper backwash can also lead to accumulation
of materials on the surface of the filter that can form mudballs and cracks, which can allow solids to pass
through the filter. A surface wash or air scour may also be helpful to prevent accumulation of mudballs
or grease. Surface wash or air scour is also helpful for traveling bridge filters. Without surface wash
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traveling bridge filters are limited to an influent TSS concentration of 40 to 50 mg/L (WEF and ASCE,
1998).
If membrane filters are used, fouling can be an important consideration. Cellulose acetate
membranes can be damaged by biological activity. Disinfection is often used to prevent biological
fouling of the membranes. Some membrane materials such as polyacramides, however, can be
damaged by chlorine. This can be avoided by using an alternative disinfectant, a different membrane
material, or by de-chlorination. Lowering the pH can help to prevent mineral fouling of nanofiltration or
reverse osmosis membranes. Besides pre-treatment, chemical cleaning of the membranes may also be
required periodically. Monitoring of effluent quality and pressure differential can be important to help
identify membrane fouling or failure.
7.3 Ongoing Research and Emerging Technologies
The use of membranes as tertiary filtration is an area that has recently expanded. Research
continues on various membrane configurations along with topics such as pre-treatment, membrane
cleaning, and removal of emerging contaminants. Fuzzy filters are also an innovative technology that is
beginning to be established in the wastewater community with several full scale projects. Other
research has focused on enhancements to existing technology. For example, the Blue-Pro system
combines continuous backwashing filters, a well known technology, with a hydrous ferric oxide coating
and ferric salt addition to remove phosphorus by adsorption as well as filtration.
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8. Mathematical Modeling
8.1 The Need for Models
WWTPs are complex systems that depend on numerous biological, chemical, and physical
processes to achieve effluent goals. Because of the complex behavior of the processes and the
variability in wastewater characteristics, biological populations, and plant design, it is not always
possible to predict how changing any one variable will affect the effluent quality. Plant designs that
work for one influent wastewater and climate may not perform well in different conditions. Pilot scale
or full scale trials can help to determine the effect of various parameters, but costs and time to cover all
possibilities may be prohibitive. Therefore, models fill an important need by enabling simulation of a
process and estimating the impact that changing parameters will have on the treatment effectiveness.
Models can be used for a number of purposes including the design of new WWTPs, the design of
retrofits or upgrades to existing plants, determining how changes in operations may affect effluent
concentrations of permitted contaminants, determining how plants will respond to changes in influent
quality or flow, and for training operators. Not all models can achieve all of these purposes, so models
should be selected with the desired use in mind.
There is some disagreement in the literature in the use of the term model. Some references use
the term to refer to sets of mathematical equations that characterize a process, other references use
model to refer to the computer program used to solve these equations. This chapter will use the former
and will use the term "simulator" to describe the computer program.
8.2 Overview of Available Models
Models are sets of equations, generally based on theory and grounded in empirical data, that
represent a wastewater treatment process. Each unit process is represented by its own model. Model
equations for processes such as clarification and settling are well known and fairly simple. Modeling
biological wastewater processes such as activated sludge, however, is much more complicated. The
primary set of models for activated sludge processes has been compiled by the International Water
Association (IWA). The first model was developed in 1986 and was called the activated sludge model
(ASM). Later known as ASM1, this model was able to model the biological oxidation of carbon,
nitrification, and de-nitrification.
Although the ASM model gained widespread use among both academia and industry, it had
limitations. For example the model assumed constant temperature and pH, did not include EBPR, and
the biological reactions did not depend on the carbon source. In order to improve the model, IWA
developed four other ASM models; ASM2, ASM2d, ASM3, and ASM3 with BioP. ASM2 and ASM2d were
intended to add EBPR. The ASM3 models were intended to deal with limitations such as the
independence of the ASM1 model of temperature and carbon source. In addition, other models were
developed to seek to improve upon the ASM model. The metabolic biological phosphorus model of the
Delft University of Technology (TUDP) was developed to fully account for the metabolism occurring in
PAOs during EBPR. Barker and Dold (1997) developed a model (B&D) to include different rates of
growth depending on the carbon source.
Table 8-1 lists each mathematical model for wastewater treatment, the processes it can
simulate, and the reference where the model equations along with its limitations and valid range of
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operating parameters can be found. When selecting models, the processes required and the range of
normal operating parameters for the plant should be considered and compared to the available models.
For example, if chemical phosphorus removal is to be used in a plant, the plant is limited to using either
the ASM2 or ASM2d models. Each model also has a range of temperatures and pH over which it is valid.
Table 8-1 Available Activated Sludge Models
Model Name
ASM1
ASM2
ASM2d
ASMS
ASMS w/ BioP
TUDP
B&D
Wastewater Treatment Unit Processes
Carbon oxidation, nitrification, de-nitrification
Carbon oxidation, nitrification, de-nitrification, enhanced
biological phosphorus removal, fermentation, chemical
phosphorus removal
Carbon oxidation, nitrification, de-nitrification, enhanced
biological phosphorus removal, fermentation, chemical
phosphorus removal
Carbon oxidation, nitrification, de-nitrification
Carbon oxidation, nitrification, de-nitrification, enhanced
biological phosphorus removal
Carbon oxidation, nitrification, de-nitrification, enhanced
biological phosphorus removal, fermentation
Carbon oxidation, nitrification, de-nitrification, enhanced
biological phosphorus removal, fermentation
Reference
Henzeetal. 1987
Henzeetal. 1995
Henzeetal. 1999
Henzeetal. 1999
Reiger et al. 2001
Brdjanovic et al. 2000
Barker and Dold 1997
Sources: WERF 2003, Gernaey et al. 2004
Fixed film processes are gaining popularity in nutrient removal. The modeling of these
processes is more complicated because of the added dimension of diffusion to and from the biofilm. An
equivalent to the ASM models does not exist for fixed film processes. IWA, however, did publish a
reference in 2006 (Wanner et al., 2006) listing all of the model equations available for fixed film
processes. These models fall into three categories: semi-empirical models, 1-dimensional models, and
2-dimensional models. 3-dimensional models exist, but do not yield significantly better results for the
added computation time. Semi-empirical models can be solved analytically, run the fastest, and require
more assumptions on the limiting factors in the reactor and biofilm. 1-dimensional and 2-dimensional
models must be solved numerically. Semi-empirical values for kinetic constants, carbon removal,
nitrification, de-nitrification, and biomass growth can increase run speeds. The reference also discusses
methods for solving the sets of equations governing fixed film processes. Solution methods vary from
analytical solutions for more simplified sets of equations to numerical solutions for more complex model
equations. Many simulators now can accommodate these new equations as well.
There are numerous simulators available that run combinations of the various models.
Simulators typically have a graphical interface which allows the user to specify the unit processes
included in the plant. Most simulators allow the selection from a number of models appropriate for the
unit processes to be represented. Different simulators run different sets of models so the selection of
the right simulator is important. In addition to selecting the required processes, the user also sets the
flow rates including recycle streams and the influent wastewater characteristics. Parameters such as
kinetic constants for biological growth and stoichiometric constants for reactions can be specified, or the
user can select default values. Given the process layout, the input parameters, and the selected model,
the simulator solves the system of equations to predict the wastewater characteristics throughout the
plant.
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Simulators currently available include: GPS-X, EFOR, STOAT, BioWin, ASIM, SIMBA, WEST,
AQUASIM, and AQUIFAS. Simulators vary in the models they can run, the degree of skill required to run
the simulator, the user interface, and the cost. Some models allow a high degree of customization, but
will require greater user knowledge of the underlying models. Others are more user friendly for less
skilled users but in such cases the limitations of the models may not be as obvious to the user.
8.3 Model Inputs
Model parameters are the input variables entered by the user and fall into three categories:
• Parameters defining reactions within the process
• Wastewater characteristics
• Process inputs
Parameters defining the reactions are selected when the user selects the model. In some cases
the user may wish to modify a model default value. Process inputs include the flows for each stream,
recycle flows, and the size of each process. These inputs are generally design parameters or are known
values in the case of an existing plant. Wastewater characteristics include influent BOD or COD,
nitrogen, and phosphorus along with temperature and pH. Most models ask for detailed breakdowns of
these parameters. COD is usually broken down into slowly biodegradable, readily biodegradable,
particulate unbiodegradable, and soluble unbiodegradable. Nitrogen is typically broken into ammonia,
soluble unbiodegradable TKN, biodegradable TKN, and particulate unbiodegradable TKN. For models
with EBPR, orthophosphate, soluble unbiodegradable phosphorus, organically bound biodegradable
phosphorus, and particulate unbiodegradable phosphorus fractions are required. Some models require
the wastewater characteristics as concentrations while others require both total concentration and
fractions.
Additional sampling and laboratory analysis to characterize the influent wastewater is needed to
fulfill model input requirements. At a minimum, the fractions of each of the wastewater components
must be determined. It may also be beneficial to determine some reaction constants through
laboratory experiments with the biomass to be used. For example, the nitrifier growth rate frequently
varies from plant to plant and measuring it directly may yield better design results.
8.4 Model Calibration
Because the models are not exact descriptions of the process and because measurement of
parameters is not exact, every model will need to be calibrated. A model is calibrated by running the
simulator and comparing predicted values of wastewater components with measured values throughout
the plant, or through a similar plant. Model parameters are then adjusted to obtain the best fit
between the simulator results and the plant data. In general the parameters for which the input values
are the least certain are the ones that are adjusted. Parameters that are commonly adjusted include the
sludge waste rate or nitrifier growth rate. Parameters are not adjusted unless they significantly improve
the model fit and the resulting parameter is still within realistic bounds. The parameters are adjusted so
the model is valid over the widest range possible rather than trying to get as close a fit as possible for a
single set of circumstances.
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9. Nutrient Removal for Small Communities and Decentralized Wastewater
Treatment Systems
Approximately 25 percent of the U.S. population is served by onsite septic or decentralized
systems. Onsite septic systems treat and dispose of effluent on the same property that produces the
wastewater, whereas decentralized treatment refers to onsite or cluster systems that are used to treat
and dispose of relatively small volumes of wastewater, generally from dwellings and businesses that are
located relatively close together. In many cases, wastewater from several homes is pretreated onsite by
individual septic tanks before being transported through alternative sewers to an offsite decentralized
treatment unit that is relatively simple to operate and maintain. The remaining 75 percent of the
population is served by centralized wastewater treatment facilities, which collect and treat large
volumes of wastewater.
There is, in fact, a growing movement toward decentralized or clustered wastewater treatment
systems to reduce cost, to provide groundwater recharge near the source, and for speed and ease in
siting since they are generally located underground. The use of residential cluster development is
gaining in popularity across the U.S. as a means to permanently protect open space, preserve
agricultural land, and protect wildlife habitat (Mega et al., 1998). As part of these developments,
wastewater systems such as community drainfields, irrigation systems, and package plants are being
installed to reduce infrastructure investment and minimize adverse environmental impacts. Additional
alternatives that include aerobic tanks, sand filters, and constructed wetlands can be used to reduce
nutrient pollution; particularly in sensitive coastal areas or over sensitive, unconfined aquifers used for
drinking water (Anderson and Gustafson, 1998).
9.1 Phosphorus Removal
Few phosphorus removal processes are well developed for onsite wastewater systems
application (USEPA, 2008e). The controlled addition of chemicals such as aluminum, iron, and calcium
compounds with subsequent flocculation and sedimentation has had only limited success because of
inadequate operation and maintenance of mechanical equipment and excessive sludge production.
Most notable successes have come with special filter materials that are naturally high in their
concentration of the above chemicals, but their service lives are finite. Studies of high-iron sands and
high-aluminum muds indicate that 50 to 95 percent of the phosphorus can be removed. However, the
life of these systems has yet to be determined, after which the filter media will have to be removed and
replaced. Use of supplemental iron powder mixed with natural sands is also being researched. Aside
from specialized filter media, the most likely phosphorus-reduction systems are iron-rich intermittent
sand filter (ISF) media and SBRs. These are discussed in more detail below.
9.2 Nitrogen Removal
Processes that remove 25 to 50 percent of total nitrogen include aerobic biological systems and
media filters, especially recirculating filters (USEPA, 2008f). The vast majority of on-site and cluster
nitrogen-removal systems employ nitrification and denitrification biological reactions. Most notable of
these are recirculating sand filters (RSFs) with enhanced anoxic modifications, SBRs, and an array of
aerobic nitrification processes combined with an anoxic/anaerobic process to perform denitrification.
Some of the combinations are proprietary. A few recently developed highly instrumented systems that
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utilize membrane solids separation following biological nitrification and denitrification are capable of
removing total nitrogen down to very low concentrations (i.e. 3-4 mg/L TN).
Nitrogen removal systems generally are located last in the treatment train prior to subsurface
wastewater infiltration system (SWIS) disposal or surface water disposal, in which case a disinfection
step is typically required. Usually, the minimum total nitrogen standard that can be regularly met is
about 10 mg/L. These technologies can be either above ground or below ground.
9.3 Nitrogen and Phosphorus Removal Technologies
9.3.1 Introduction
This section provides information on a number of different technologies that can reduce
nitrogen and phosphorus levels. The actual technology selected will be site-specific and dependent on
many factors including soil conditions, influent water quality, required effluent levels, disposal options,
availability of land, cost, etc. In some cases, a combination of technologies may be necessary to
effectively remove all the contaminants of concern. Small system owners and operators should work
closely with their state onsite and decentralized program staff as well as engineers to ensure that the
technologies selected will work effectively in combination to achieve the effluent goals.
9.3.2 Nutrient Removal Technologies
Fixed-film systems - Aerobic/anaerobic trickling filter package plant
Fixed-film systems (FFSs) are biological treatment processes that employ a medium such as rock,
plastic, wood, or other natural or synthetic solid material that will support biomass on its surface and
within its porous structure (USEPA, 2008c). Trickling filter FFSs are typically constructed as beds of
media through which wastewater flows. Oxygen is normally provided by natural or forced ventilation.
Commercial on-site systems use synthetic media and receive wastewater from overlying sprayheads for
aerobic treatment and nitrification. Nitrified effluent returns to the anoxic zone to mix with either septic
tank contents or incoming septic tank effluent for denitrification. A portion of the denitrified effluent is
discharged for disposal or further treatment. Aerobic tanks are available in residential or small-
community sizes. Typical trickling filters systems currently available are capable of producing effluent
BOD and TSS concentrations of 5 to 40 mg/L. Nitrogen removal typically varies from 0 to 35 percent
although removal percentages as high as 65% have been demonstrated through USEPA's Environmental
Technology Verification (ETV) program. Phosphorus removal is typically 10 to 15 percent. Higher
removal occurs at low loading rates in warm climates. Systems can be configured for single-pass use
where the treated water is applied to the trickling filter once before being disposed of, or for multi-pass
use where a portion of the treated water is cycled back to the septic tank and re-treated via a closed-
loop. Multi-pass systems result in higher treatment quality and assist in removing Total Nitrogen (TN)
levels by promoting nitrification in the aerobic media bed and denitrification in the anaerobic septic
tank.
Factors affecting performance include influent wastewater characteristics, hydraulic and organic
loading, medium type, maintenance of optimal DO levels, and recirculation rates.
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Sequencing batch reactor (SBR)
The SBR process is a sequential suspended growth (activated sludge) process in which all major
steps occur in the same tank in sequential order (USEPA, 2008d). The SBR system is typically found in
packaged configurations for onsite and small community or cluster applications. The major components
of the package include the batch tank, aerator, mixer, decanter device, process control system (including
timers), pumps, piping, and appurtenances. Aeration may be provided by diffused air or mechanical
devices. SBRs are often sized to provide mixing as well and are operated by the process control timers.
Mechanical aerators have the added value of potential operation as mixers or aerators. The decanter is
a critical element in the process. Several decanter configurations are available, including fixed and
floating units. At least one commercial package employs a thermal processing step for the excess sludge
produced and wasted during the "idle" step. The key to the SBR process is the control system, which
consists of a combination of level sensors, timers, and microprocessors which can be configured to meet
the needs of the system.
SBRs can be designed and operated to enhance removal of nitrogen, phosphorus, and ammonia,
in addition to removing TSS and BOD. Package plant SBRs are suitable for areas with little land, stringent
treatment requirements, and small wastewater flows such as RV parks or mobile homes, campgrounds,
construction sites, rural schools, hotels, and other small applications. These systems are also useful for
treating pharmaceutical, brewery, dairy, pulp and paper, and chemical wastes (USEPA, 2000d).
Intermittent sand filters (ISF)
ISF is used to describe a variety of packed-bed filters of sand or other granular materials
available on the market (USEPA, 2008g). Sand filters provide advanced secondary treatment of settled
wastewater or septic tank effluent. They consist of a lined (e.g., impervious PVC liner on sand bedding)
excavation or structure filled with uniform washed sand that is placed over an underdrain system. The
wastewater is directed onto the surface of the sand through a distribution network and allowed to
percolate through the sand to the underdrain system. The underdrain system collects the filter effluent
for further processing or discharge.
Sand filters are aerobic, fixed-film bioreactors. Bioslimes from the growth of microorganisms
develop as films on the sand particle surfaces. The microorganisms in the slimes capture soluble and
colloidal waste materials in the wastewater as it percolates over the sand surfaces. The captured
materials are metabolized into new cell mass or degraded under aerobic conditions to carbon dioxide
and water. Most biochemical treatment occurs within approximately 6 inches of the filter surface.
Other treatment mechanisms that occur in sand filters include physical processes, such as
straining and sedimentation, to remove suspended solids within the pores of the media. Most
suspended solids are strained out at the filter surface. Chemical adsorption can occur throughout the
media bed. Adsorption sites in the media are usually limited, however. The capacity of the media to
retain ions depends on the target constituent, the pH, and the mineralogy of the media. Phosphorous is
one element of concern in wastewater that can be removed in this manner, but the number of available
adsorption sites is limited by the characteristics of the media.
Sand filters can be used for a broad range of applications, including single-family residences,
large commercial establishments, and small communities. Sand filters are frequently used to pretreat
septic tank effluent prior to subsurface infiltration onsite where the soil has insufficient unsaturated
depth above ground water or bedrock to achieve adequate treatment. They are also used to meet
water quality requirements (with the possible exception of fecal coliform removal) before direct
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discharge to surface water. Sand filters are used primarily to treat domestic wastewater, but they have
been used successfully in treatment trains to treat wastewaters high in organic materials such as those
from restaurants and supermarkets. Single-pass ISF filters are most frequently used for smaller
applications and sites where nitrogen removal is not required. However, they can be combined with
anoxic processes to significantly increase nitrogen removal.
Recirculating sand filters (RSF)
Recirculating filters using sand, gravel, or other media provide advanced secondary treatment of
settled wastewater or septic tank effluent (USEPA, 2008h). They consist of a lined (e.g., impervious PVC
liner on sand bedding) excavation or structure filled with uniform washed sand that is placed over an
underdrain system. The wastewater is directed onto the surface of the sand through a distribution
network and allowed to percolate through the sand to the underdrain system. The underdrain system
collects and recycles the filter effluent to the recirculation tank for further processing or discharge.
The basic components of recirculating filters include a recirculation/dosing tank, pump and
controls, distribution network, filter bed with an underdrain system, and a return line. The return line or
the underdrain must split the flow to recycle a portion of the filtrate to the recirculation/dosing tank. A
small volume of wastewater and filtrate is dosed to the filter surface on a timed cycle 1 to 3 times per
hour. Recirculation ratios are typically between 3:1 and 5:1. In the recirculation tank, the returned
aerobic filtrate mixes with the anaerobic septic tank effluent before being reapplied to the filter.
RSFs can be used for a broad range of applications, including single-family residences, large
commercial establishments, and small communities. They produce a high quality effluent with
approximately 85 to 95 percent BOD and TSS removal. In addition, almost complete nitrification is
achieved. Denitrification also has been shown to occur in RSFs. Depending on modifications in design
and operation, 50 percent or more of applied nitrogen can be removed (USEPA, 1999). To enhance this
capability, they can be combined with a greater supply of biodegradable organic carbon, time, and
mixing than is normally available from the conventional recirculation tank.
Natural Systems
The natural systems described here include constructed wetlands and floating aquatic plant
treatment systems. Wetland systems are typically described in terms of the position of the water
surface and/or the type of vegetation grown. Most natural wetlands are free water surface (FWS)
systems where the water surface is exposed to the atmosphere; these include bogs (primary vegetation
mosses), swamps (primary vegetation trees), and marshes (primary vegetation grasses and emergent
macrophytes) (USEPA, 2000e). subsurface flow (SF) wetlands are specifically designed to treat or polish
wastewater and are typically constructed as a bed or channel containing appropriate media.
Constructed wetlands treat wastewater by bacterial decomposition, settling, and filtering. As in
tank designs, bacteria break down organic matter in the wastewater, aerobically, anoxically and
anaerobically. Oxygen for aerobic decomposition is supplied by the plants growing in the wetland.
Solids are filtered and finally settle out of the wastewater within the wetland. After about two weeks in
the wetland, effluent is usually discharged by gravity to an unlined wetland bed. If these systems
discharge effluent to surface ditches, they require a NPDES permit.
The submerged plant roots do provide substrate for microbial processes. However, the amount
of oxygen that emergent macrophytes can transmit from the leaves to their roots is negligible compared
to the oxygen demand of wastewater. Therefore subsurface flow wetlands are devoid of oxygen. The
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lack of oxygen in these subsurface flow systems means that ammonia oxidation via biological
nitrification will not occur without the use of an additional unit process, such as a gravel trickling filter
for nitrification of the wastewater ammonia. Vertical flow wetland beds are a modification of
subsurface flow wetlands which contain gravel or coarse sand and are loaded intermittently at the top
surface. Unlike ammonia oxidation, nitrate removal in a subsurface flow wetland can be rapid and
effective because the anoxic conditions and carbon sources necessary to support the treatment
reactions occur naturally in these systems.
FWS wetlands with long detention times can remove minor amounts of phosphorus through
plant uptake, adsorption, complexation, and precipitation. However, removal via plant uptake is limited
to phosphorus retained in plant litter that is buried by sediments before plant decomposition occurs (i.e.
peat building process). Phosphorus removal is typically greater in the first year or two because of soil
absorption and rapidly expanding vegetation but decreases when the system reaches equilibrium, and
unburied plant litter releases phosphorus back into the water as it decomposes. Phosphorus removal is
also possible with the use of an addition process, such as chemical addition and mixing prior to a final
deep settling pond.
Aquatic systems using duckweed have been used for a number of years to treat wastewater for
various purposes (WEF, 2001). Duckweed (Lemna spp.) are floating macrophytes. Duckweed fronds can
double their mass in two days under ideal conditions of nutrient availability, sunlight, and temperature.
Although duckweed can be found in most regions, the rate of growth is optimal at 20 to 30 °C and they
grow best in a pH range of 3.5 to 8.5. Duckweed can grow about six months per year in most U.S.
climates. High levels of BOD and TSS removal have been observed from duckweed systems. To achieve
secondary treatment most duckweed systems are coupled with either facultative or aerated ponds.
Nitrogen is removed by plant uptake and harvesting, by denitrification, or a combination of the two.
Typically less than 1 mg/L of phosphorus can be removed by plant uptake and harvest. If significant
phosphorus removal is required, chemical precipitation with alum, ferric chloride, or other chemicals
used in a separate treatment step is necessary. The major disadvantage of duckweed systems is the
large amount of biomass produced by the rapidly growing plants, which creates a solids handling
requirement similar to handling sludge at an aerobic wastewater treatment facility.
Proprietary Filters/Improved and Emerging Technologies
A number of companies have developed proprietary nitrogen and phosphorus removal
technologies that can be used at centralized wastewater treatment facilities as well as at onsite,
decentralized systems. This section provides a general description of some of these technologies
without mentioning specific trade names. Additional information on proprietary and emerging
technologies is available from the Wastewater Virtual Trade Show hosted by EPA New England's Center
for Environmental Industry and Technology (CEIT), available online at
http://www.epa.gov/ne/assistance/ceitts/wastewater/index.html.
In one onsite technology system, the household's greywater (wash water) and blackwater (toilet
and kitchen sink wastes) are plumbed separately and flow to separate septic tanks. The system consists
of two septic tanks, a filter, and a conventional leaching facility, all of which are located below the
ground surface. In some situations the system may be passive, requiring no pumps or other moving
mechanical parts (unless finished effluent must be pumped up to an elevated leaching field to achieve
adequate separation to groundwater). The blackwater flows from the septic tank to a proprietary
aerobic filter where it is nitrified and then flows to the anaerobic greywater septic tank for
denitrification. The filter itself is composed of several layers of in-drains which are overlain by layers of
sand and filter cloth. The in-drains are composed of proprietary material and provide air to the filter so
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that it remains aerobic. The in-drain media and the sand support the growth of nitrifying bacteria. As
the effluent trickles through the filter, nitrification occurs. The effluent is collected by a drain at the
bottom of the filter from which it flows into the greywater septic tank. The greywater septic tank is
anaerobic and the greywater provides a rich source of carbon which supports the growth of denitrifying
bacteria. Denitrification of the effluent is accomplished by passive mixing of the filter-treated blackwater
with the greywater in this septic tank. After treatment in this tank, finished effluent flows to the
leaching facility for disposal. Initial data indicate that the system is capable of producing finished
effluent with total nitrogen content of 19 mg/L Limited data show that the system may also be capable
of approximately 50% removal of phosphorus. The system has also been modified for commercial use.
A centralized system for nitrogen removal has been used at full scale in the Netherlands
although it is just in pilot system phase here in the U.S. This single reactor system for nitrogen removal
is based on two core concepts: 1) that at temperatures between 30-40°C, the growth rate of ammonia
oxidizing bacteria is greater than the growth of nitrate oxidizing bacteria; and 2) denitrifying bacteria are
capable of anoxic conversions of nitrite to nitrogen gas. The reactor converts ammonia mainly to nitrite
by oxidation at a minimal sludge retention time. The nitrite is then anoxically converted to nitrogen gas.
A passive wastewater nitrogen removal system has been developed at the University of
Waterloo, Ontario and has been tested in several facilities in the U.S. This system, which can reduce
wastewater nitrogen levels to 3 mg/L, has been demonstrated to be effective for individual properties as
well as cluster and centralized wastewater systems. The proprietary patented nitrate-reactive media
that converts nitrate to inert nitrogen gas (denitrification) is contained in a prefabricated tank or, for
larger installations, in an engineered excavation. Nitrate contaminated wastewater is gravitationally fed
through the treatment module. For septic tank applications, an oxidative pre-treatment step is required
to convert ammonium (NH4+) to nitrate (NO3~) before the filter can perform the reductive denitrification
step. Pre-treatment can be achieved with any of the existing oxidative technologies commonly used in
wastewater treatment. The nitrate-free effluent from the filter is then discharged to a conventional tile
bed or receiving water body. The filters have also been used as a permeable reactive barrier to remove
groundwater nutrients close to receiving water bodies.
A passive phosphorus wastewater removal system has also been developed at the University of
Waterloo that incorporates proprietary media and a filter following primary treatment/septic tanks and
prior to soil dispersal. This system has been demonstrated to reduce phosphorus effluent by 94 percent,
nitrogen by 54 percent and BOD by 86 percent at a test center in Massachusetts. The process
incorporates reductive iron dissolution and mineralization of phosphorus.
Another manufacturer utilizes attached growth airlift reactor technology with a fully open, fully
protected biomass carrier with aeration and mixing design to provide high BOD and total nitrogen
removal capability. This is a fixed biofilm moving bed process, utilizing suspended biomass carriers with
high specific surface area for biofilm growth, along with carefully designed reactor hydraulics. This dual
action denitrification and effluent filtration solution technology has been designed for small and
medium sized WWTPs as well as other industrial applications such as food and beverage, pulp and
paper, oil and gas, bio fuel, chemicals, Pharmaceuticals, reuse, and aquaculture.
A variety of membranes and bioreactor technologies, many using improved fixed-film processes,
are also being developed for maximum exposure of waste streams to activated sludge that digests waste
from plant production effluent streams. Since the membranes can retain activated sludge at high
concentrations, BOD and nitrogen can be removed more efficiently. Many of these can be integrated as
an upgrade/retrofit into existing treatment systems or used as package plants in smaller installations
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such as residential and commercial development projects, industrial applications, military bases, and
office parks.
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10. Sustainable Nutrient Recovery
While the U.S. is primarily addressing nutrient removal concerns through development of WQSs
and treatment at centralized wastewater facilities, a number of European countries including
Switzerland, Sweden, and the Netherlands are conducting research on innovative sustainable nutrient
recovery systems. The concept behind these new technologies is to separate and treat toilet waste
before it leaves the home or building and mixes with the larger waste stream to be carried to WWTPs.
Recent studies have shown that about 80 percent of the nitrogen and 50 percent of the phosphorus in
wastewater are derived from urine although urine makes up only 1 percent of the volume of
wastewater (Larsen and Leinert, 2007). Separating the urine from wastewater could offer various
advantages: WWTPs could be built on a smaller scale, water bodies will be better protected from
nitrogen and phosphorus pollution, nutrients could be recycled for agricultural use, and various
constituents of concern including hormones and pharmaceutical compounds could be removed before
being mixed with wastewater and released to the environment. A major benefit would be reduced
energy consumption at WWTPs as a result of reduced treatment requirements for nitrogen. Also,
separating 50 to 60 percent of urine could reduce in-plant nitrogen gas discharges and result in fewer
impurities in methane captured from sludge digestion.
Organizations such as the Swiss Federal Institute of Aquatic Science and Technology (Eawag) are
currently experimenting with the development and application of "NoMix technology" to separate urine
from solid waste at the toilet bowl. While similar in size and shape to current toilets, this new
technology has two waste pipes - a small front one that collects and diverts urine into a storage tank,
and a larger rear waste pipe that operates like a standard toilet. The first of these toilets were installed
in two "eco-villages" in Sweden in 1994 and since then have spread to other locations throughout the
country and to Denmark, the Netherlands, and Switzerland. The concept is now taking hold in Austria
and Germany. While the pollutant-free urine, or "urevit," can be spray-applied directly onto agricultural
fields; in the Netherlands, a company called Grontmij trucks stored urine to a special treatment plant
where the phosphate is precipitated out as a mineral called struvite and used as a fertilizer. Novaquatis,
a branch of Eawag is experimenting with extracting nitrogen and potassium from urine that can be
sprayed directly onto crops. Eawag is also experimenting with a pilot decentralized basement sewage
plant where domestic wastewater is treated in a MBR so it can be reused for flushing the toilets or
watering the garden and the sewage sludge is composted. While still experimental, some of these
technologies may have practical future applications if widely applicable low-cost solutions can be found
for urine transport, or stable and cost-effective technologies can be developed for decentralized
treatment. While studies of consumer attitudes and acceptance appear to be positive, technological
improvements are still needed to prevent clogging in pipes, to identify best treatment options that can
be applied in practice; and to identify how and where to convert urine to fertilizer.
Sustainability concerns are also driving the wastewater treatment industry to start looking at
sludge as a renewable resource. Historically, agricultural use has been the traditional approach for
disposal of municipal sludge due to its high nutrient content for fertilizing crops, and its low cost
approach. As scientific advances detect smaller and smaller quantities of contaminants (i.e., heavy
metals, pathogenic microorganisms, Pharmaceuticals, and personal care products), the public, farming
organizations, and the food industry are raising concerns about continuing this practice. As noted
above, researchers are discovering that valuable products can be generated from sewage treatment
byproducts such as energy extracted from anaerobic digestion, construction materials such as bricks,
and nutrients such as phosphorus that can be extracted from sludge and used as fertilizer.
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In February 2008, the non-profit Global Water Research Coalition, an international water
research alliance formed by 12 world-leading research organizations, released a report titled, State of
Science Report: Energy and Resource Recovery from Sludge (Kalogo and Monteith, 2008). The report
focuses on:
• The international situation of energy and resource recovery from sludge
• How the use of different sludge treatment processes affects the possibility of recovering
energy and/or materials from the residual sludge
• The influence of market and regulatory drivers on the fate of the sludge end-product
• The feasibility of energy and resource recovery from sludge
• The social, economic, and environmental performance (triple bottom line orTBL
assessment) of current alternatives technologies
• Four market drivers are identified and discussed including:
Sustainability and environmental concerns, such as the threat of soil pollution, global
warming and resource depletion
Rising energy costs and the need of more electricity and heat to operate the plants
Requirements for high quality of resources for industrial applications, such as calcium
phosphate for the phosphate industry
Regulation as a factor stimulating the development of new technologies
In the report, energy recovery technologies are classified into sludge-to-biogas processes,
sludge-to-syngas processes, sludge-to-oil processes, and sludge-to-liquid processes. The technologies
available for resource recovery discussed in the report include those to recover phosphorus, building
materials, nitrogen, and volatile acids. The report, which covers both established as well as emerging
technologies, will be used as the basis for development of the coalition's future strategic research plan
on energy and recovery from sludge. As a technical resource, it provides a valuable overview of sludge
disposal practices in various countries such as the U.S., the Netherlands, the United Kingdom, Germany,
Sweden, Japan, and China; and presents a number of treatment processes for resource recovery.
Other groups have looked at recovering phosphorus from the supernatant from anaerobic
digestion. Several different processes have been proposed that rely on precipitation of the phosphorus
as either struvite or calcium phosphate. Work is underway on projects in Italy, Germany, the
Netherlands, and Canada (SCOPE, 2004).
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11. Co-Removal of Emerging Contaminants
This chapter provides a brief background on emerging contaminants and key findings from
studies on the co-removal of emerging contaminants by nutrient removal technologies.
11.1 Background on Emerging Contaminants
The term "emerging contaminants" refers broadly to those synthetic or naturally occurring
chemicals, or to any microbiological organisms, that have not been commonly monitored in the
environment but which are of increasing concern because of their known or suspected adverse
ecological or human health effects. Emerging contaminants can fall into a wide range of groups defined
by their effects, uses, or by their key chemical or microbiological characteristics. Two groups of
emerging contaminants that are of particular interest and concern at present are endocrine disrupting
chemicals (EDCs) and pharmaceutical and personal care products (PPCPs). These compounds are found
in the environment, often as a result of human activities.
EDCs may interfere with the endocrine systems by damaging hormone-producing tissues,
changing the processes by which hormones are made or metabolized, or mimicking hormones. In
addition to natural and synthetic forms of human hormones that are released into the environment,
there are a multitude of synthetic organic compounds that are able to disrupt the endocrine system.
Public concern about EDCs in the environment has been rapidly increasing since the 1990s when
researchers reported unusual sexual characteristics in wildlife. A report by the USGS, found that fish in
many streams had atypical ratios of male and female sex hormones (Goodbred et al., 1997). In England,
researchers found that male trout kept in cages near WWTP outfalls were developing eggs on their
testes and had increased levels of the protein that is responsible for egg production (vitellogenin)
(Sumpter, 1995; Kaiser, 1996). Follow-up laboratory studies showed that synthetic forms of estrogen
(17ct-ethynylestradiol (EE2)) could increase vitellogenin production in fish at levels as low as 1-10 ng/L,
with positive responses seen down to the 0.1-0.5 ng/L level (Purdom et al., 1994).
Human estrogens have the ability to alter sexual characteristics of aquatic species at trace
concentrations as low as 1 ng/L (Purdom et al., 1994). WWTP effluents have been identified as a
primary source for EDCs in the environment, with the bulk of their endocrine disrupting activity resulting
from human estrogen compounds (Desbrow et al., 1998, Snyder et al., 2001) (Table 11-1). The synthetic
estrogen, EE2, and the natural estrogens, estrone (EJ and l?p-estradiol (E2), are the greatest
contributors to endocrine disrupting activity in WWTP effluent (Johnson et al., 2001) with EE2 showing
the greatest recalcitrance in WWTPs (Joss et al., 2004). Influent concentrations range from below
detection to 70 ng/L for EE2, 670 ng/L for EI and 150 ng/L for E2 (Vethaak et al., 2005, Clara et al.,
2005b).
Other EDCs include tributyl tin, which was previously used in paints to prevent marine
organisms from sticking to ships, nonylphenol (a surfactant), and bisphenol A (platicizer and
preservative).
PPCPs encompass a wide variety of products that are used by individuals for personal health or
cosmetic reasons, and also include certain agricultural and veterinary medicine products. PPCPs
comprise a diverse collection of thousands of chemical substances, including prescription and over-the-
counter therapeutic drugs, veterinary drugs, fragrances, sun-screen products, vitamins, and cosmetics.
Many of these products, notably the Pharmaceuticals for human or animal use, are specifically designed
to be biologically active, and some PPCPs may also fall into the category of EDCs described previously.
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Table 11-1. Estrogens of Concern
Name
Chemical Structure
Name
Chemical Structure
Ex
Estrone
E2
17p-estradiol
E3
Estriol
JO
EE2
17ot-
ethynylestradiol
Currently, municipal sewage treatment plants are engineered to remove conventional pollutants
such as solids and biodegradable organic material but are not specifically designed for PPCP removal or
for other unregulated contaminants. Wastewater treatment commonly consists of primary settling
followed by biological treatment, secondary settling, and disinfection. This treatment can remove more
than 90 percent of many of the most commonly known or suspected EDCs found in wastewater influent;
however, low concentrations of some suspected EDCs may remain in the wastewater treatment sludge
or effluent (WERF, 2005). As discussed in the next section, studies have shown enhanced nutrient
removal technologies to be effective in removing low concentrations of some emerging contaminants.
11.2 Removal of Emerging Contaminants by Nutrient Removal Technologies
Several studies have examined the effectiveness of current wastewater treatment technologies
in the removal of emerging contaminants. Some of these studies are discussed below and their major
findings are organized under three subsections: role of activated sludge SRT in removal efficiency, role
of nitrifying bacteria in biodegradation, and use of RO to improve removal efficiencies. Details regarding
the study design, such as evaluated treatments and contaminants, and a summary of major study
findings are provided by author in Table 11-2 at the end of this chapter. The significant findings are also
presented as follows:
• Removal efficiencies were enhanced for several investigated contaminants at longer SRTs, with
critical SRTs for some beyond which removal rates did not improve.
• Longer SRTs allow for the establishment of slower growing bacteria (e.g., nitrifying bacteria in
activated sludge), which in turn provide a more diverse community of microorganisms with
broader physiological capabilities.
• Nitrifying bacteria may play a key role in biodegradation but the role of heterotrophic bacteria
may also play a significant role.
• Reverse osmosis has been found to effectively remove PPCPs below detection limits including
those that that were not consistently removed at longer SRTs.
One caveat regarding studies on emerging contaminants is that their concentrations in
wastewater influent are often quite low (e.g., concentrations of ng/L to u.g/L range) and may be close to
method detection limits. Therefore, small variations between measured influent and effluent
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concentrations may show large variations in apparent removal efficiencies, possibly even producing
negative calculated removals.
Role of Solids Retention Time in Removal Efficiency
The focus of several studies has been the relationship of the SRT to the removal of emerging
contaminants. In particular, many investigated whether longer SRTs would result in increased removal
efficiencies for estrogens and other categories of PPCPs. Longer activated sludge SRTs allow for the
establishment of slower growing bacteria (e.g., nitrifying bacteria in activated sludge), which in turn
provide a more diverse community of microorganisms with broader physiological capabilities.
Clara et al. (2005a), Kreuzinger et al. (2004), and Oppenheimer et al. (2007) observed enhanced
removal with increasing SRTs for most of the EDCs and Pharmaceuticals tested and found no significant
differences in removal performances between conventional activated sludge systems and MBR when
operated at similar SRTi0°c. This is likely due to the molecular weight of the study compounds, which
was smaller than the molecular weight cut-off of the ultrafiltration membranes in the MBR.
Researchers have observed similar findings for natural estrogens with higher removal
percentages at longer SRTs. Effluent concentrations for three natural estrogens were measured near
their detection limits at SRTsi0°c higher than 10 days, with their critical SRTsi0°c estimated between 5 and
10 days (Clara et al., 2005a). High removal rates of > 90 percent were also observed by Joss et al. (2004)
in a study in which they evaluated the removal of EI, E2, and EE2 under aerobic and anaerobic conditions
in WWTPs designed for nutrient removal. Joss et al. (2004) also reported that the maximum efficiency is
dependent on redox conditions, with the highest removal rate occurring during the reduction of Eito E2
under aerobic conditions. Clara et al. (2005a) cited examples where conflicting results were obtained
for EE2. Ternes et al. (1999) found no significant elimination of this compound during batch
experiments; however, Baronti et al. (2000) and Joss et al. (2004) report greater than 85 percent
removal in full-scale WWTPs.
For the Pharmaceuticals ibuprofen and bezafibrate, Clara et al. (2005a) reported more than 95
percent removal during treatment and calculated the critical value for SRTi0°c at 5 days for ibuprofen
and about 10 days for bezafibrate. Analogous removal results were obtained in several other studies
(Stumpf et al., 1998; Buser et al., 1999; Zwiener et al., 2001, as cited in Clara et al., 2005a; Oppenheimer
et al., 2007). Clara et al. (2005b) noted no or slight removal of these two Pharmaceuticals and two musk
fragrances (tonalide and galaxolide) at a WWTP with a low SRT of 1 to 2 days (this plant is a high-loaded
plant, designed to remove carbon only.) Clara et al. (2005a, 2005b) also found that the pharmaceutical
carbamazepine was not removed during wastewater treatment. In addition, these studies found
contradictory results for diclofenac (e.g., removal rates ranged from no removal to > 70 percent at SRTs
of > 10 days (Clara et al., 2005b)). Clara et al. (2005a) also cited several examples where conflicting
results were obtained for diclofenac. No significant removal was reported by Buser et al. (1999) and
Heberer (2002a); whereas, Ternes et al. (1998) observed elimination rates of up to 70 percent.
Clara et al. (2005a, 2005b) concluded that the removal potential for conventional WWTPs and
MBRs depends on the SRT. They further concluded that high removal rates can be achieved at SRTsi0°c
of more than 10 days. These parameters correspond to the design criteria for nitrogen removal in the
German Association for Water, Wastewater and Waste (ATV-DVWK, 2000) and the urban wastewater
directive of the European Community (91/271/EEC) for WWTPs in sensitive areas.
In its 2005, technical brief, "Endocrine Disrupting Compounds and Implications for Wastewater
Treatment," WERF summarized information from several studies that examined the effectiveness of
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current wastewater treatment technologies in the removal of EDCs. The classes of EDCs included:
steroids/sterols (naturally occurring, synthetic, and phytoestrogens), organohalides,
metals/organometals, alkyl phenols, polycyclic aromatic hydrocarbons (PAHs)/crude oil, and plasticizers.
Although the WERF 2005 technical brief states that in general, EDC treatment effectiveness is improved
with increased SRT, it does not provide the specific SRTs that are associated with the cited removal
rates.
Oppenheimer et al. (2007) examined the relationship of SRT to treatment removal efficiencies
for 20 PPCPs that are commonly found in the influent of U.S. treatment facilities. Many of the studies
already discussed here have been conducted primarily in Europe, were conducted at small-scale WWTPs
and bench/pilot plants under controlled conditions, and focused on estrogens and prescription
Pharmaceuticals rather than PPCPs. The Oppenheimer et al. (2007) study also noted trends regarding
the effect of HRT and pure oxygen systems compared to conventional aeration systems on PPCP
removal.
Oppenheimer et al. (2007) defined a minimum critical SRT as the minimum time needed to
consistently demonstrate greater than 80 percent removal. The results of the study showed that this
critical SRT was compound dependent but that the majority of the 20 PPCPs were consistently removed
in those treatment plants operating at SRTs of 5 to 15 days. Specifically, 9 of 12 frequently occurring
PPCPs were effectively removed through secondary treatment (e.g., ibuprofen). Conversely, six
compounds that are routinely detected in influent (i.e., detected in at least 20 percent of the influent
samples) were not well removed by secondary treatment (BHA, DEET, musk ketone, triclosan,
benzophenone, galaxolide). The results for galaxolide conflicted with those reported by Clara et al.
(2005b) who generally found high removal rates with SRTs > 10 days and Kreuzinger et al. (2004) who
reported removal at SRT between 25 to 40 days. Oppenheimer et al. (2007) found that some
compounds such as octylphenol, tri-(chloroethyl) phosphate, and triphenylphosphate were not well
removed by secondary treatment; however, these were seldom detected in the influent samples.
Based on these results, Oppenheimer et al. (2007) concluded that secondary treatment provides an
"effective first barrier" for the 20 PPCPs in the study.
Oppenheimer et al. (2007) also noted trends regarding the effect of HRT and pure oxygen
systems compared to conventional aeration systems on PPCP removal but determined that insufficient
data existed to make any definitive conclusions. When the PPCP removal performance of a high-purity
oxygenated activated sludge plant was compared to a conventional aeration system, the pure oxygen
system showed higher removal rates although its SRT was shorter than the conventional aeration plant
(i.e., 1 day versus 3 days). In addition, different HRTs operating at similar SRTs had similar removal
rates, and therefore suggested that HRT does not significantly affect removal effectiveness in the
investigated PPCPs.
Role of Nitrifying Bacteria in Biodegradation
As discussed above, longer SRTs allow for the establishment of slow-growing nitrifying bacteria
(i.e., ammonia oxidizing bacteria and nitrite-oxidizing bacteria). Several studies evaluated whether
nitrifying bacteria improve the biodegradation of certain emerging contaminants. Major findings from
some of these studies are discussed in this section.
The WERF (2005) technical brief indicated that secondary biological treatment that includes
nitrification, nutrient removal, and disinfection may remove more than 90 percent of certain steroids,
and >95 percent of alkyl phenols; whereas, secondary biological treatment without nitrification and
disinfection may decrease removal of these by more than 15 percent.
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Batt et al. (2006) investigated the role of nitrifying bacteria in activated sludge in the
biodegradation of two Pharmaceuticals, iopromide and trimethoprim. The biodegradation of these
compounds was conducted in two lab-scale bioreactors using biomass from a stage-2 activated sludge
WWTP (operated at an SRT of 49 days). In one of the bioreactors, nitrification was not inhibited (Batch-
1 reactor); in the other, nitrification was inhibited with allylthiourea (Batch-2 reactor). Monitoring was
also conducted in the WWTP and compared to results obtained from the batch reactors.
Both reactors exhibited high removal rates for iopromide; however for trimethoprim, Batch-1
showed a high removal rate of 70 percent, contrasted to the Batch-2 reactor removal rate of
approximately 25 percent with nitrification inhibited. Removal rates within the treatment plant,
however, were consistent for both Pharmaceuticals, showing significantly higher removal rate after
nitrification (approx. 60 percent for iopromide and 50 percent for trimethoprim) compared to activated
sludge treatment only ( <1 percent for both). Based on these results, Batt et al. (2006) concluded that
nitrifying bacteria have a key role in the biodegradation of Pharmaceuticals in WWTP that are operated
at higher SRTs. This conclusion is supported by Marttinen et al. (2003), who investigated the fate of
phthalates in a WWTP with nitrogen removal and observed that about one third of the removal
occurred in the nitrification/denitrification treatment phase.
Studies by Yi and Harper (2007), Khunjar et al. (2007), and others have focused on the
mechanisms of estrogen removal during nitrification. Possible mechanisms include sorption of
estrogens to solids and biotransformation within the treatment facility, especially in the presence of
nitrifying activated sludges (Khunjar et al., 2007). Ammonia oxidizing bacteria have monoxygenase
enzymes for ammonia oxidation and these enzymes have been shown previously to be nonspecific and
able to accomplish cometabolic degradation of recalcitrant organics. Cometabolic degradation is a
reasonable hypothesis for estrogen degradation because this compound is present at low ng/L
concentrations that are below those expected to support microbial growth on that compound alone.
One goal of the Yi and Harper (2007) study was to establish whether biotransformation of EE2 is
due to cometabolic activity. They conducted batch experiments using enriched cultures of autotrophic
ammonia oxiders. Their study and others (Vader et al., 2000, Shi et al., 2004, as reported in Yi and
Harper, 2007) showed a strong relationship between nitrification and EE2 removal in enriched nitrifying
cultures. Based on batch tests with and without a nitrifying bacteria inhibitor, they concluded that EE2
biotransformation can be cometabolically mediated in bioreactors that are enriched for autotrophic
nitrifiers. However, Yi and Harper (2007) noted that the heterotrophic microorganisms, if present in
activated sludge processes, may also be responsible for some micropollutant biotransformations.
Further work is needed in this area as these tests did not identify the EE2 degradation product to
confirm cometabolic degradation and the role of heterotrophs was not accounted for in some tests.
The focus of a Khunjar et al. (2007) study was to identify the role of ammonia oxidizing bacteria
compared to heterotrophic bacteria in the biotransformation of EE2. They used pure cultures of
ammonia oxidizing Nitrosomonas europaea and heterotrophic cultures that were enriched with
monooxygenase and dioxygenase enzyme systems. Nitrifying activated sludge mixed liquors were taken
from two WWTPs to seed the cultures. EE2 concentrations were 10-15 M8/L. The results of their study
showed significant sorption of EE2 to the predominantly heterotrophic culture but none to the N.
europaea culture. In addition, biotransformation of EE2 was significant in the N. europaea culture. They
observed three major EE2 metabolites at different phases of N. europaea culture growth that suggest
differential action on each byproduct by the nitrifying bacteria; however, additional work is needed to
identify these byproducts. The authors also noted that additional research is needed with continuous
flow cultivated N. europaea to determine whether these metabolites are likely to be present in nitrifying
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activated sludge. Also, N. europaea was not significantly inhibited at EE2 concentrations at or below 10
u.g EE2/L, suggesting that ammonia oxidation may not be significantly impacted by concentrations of EE2
that may be typical of those found in the environment.
Use of Reverse Osmosis to Improve Removal Efficiencies
Several studies describe the effectiveness of RO in the removal of PPCP and EDCs from
secondary wastewater effluent. Braghetta et al. (2002) calculated the removals rates that could be
achieved with a RO step following tertiary treatment for 17 PPCPs. They estimated removals to be > 90
percent for most of the selected compounds. Lower removal rates were estimated for diclofenac (55.2
to 62 percent), ketoprofen (64.3 percent), and paraxanthine (73.7 percent).
As previously discussed, the WERF (2005) technical brief evaluated RO removal rates for several
compounds. Specifically, the WERF brief cites numerous studies in which RO achieved removal rates of
90 percent or better for naturally occurring and synthetic steroids, organohalides, metals/organometals,
and alkyl phenols. In addition, Oppenheimer et al. (2007) found that RO was effective in removing all 20
investigated PPCPs below the detection limit including those that were not consistently removed at SRTs
of 30 days (i.e., galaxolide) using conventional activated sludge treatment or media filtration. Similar
findings were reported by Snyder (2003) and Everest et al. (2003) (as cited in Oppenheimer et al., 2007).
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Table 11-2. Study Design Parameters
Table 11-2. Study Design Parameters
Authors
Study Location
Study Contaminants1
Treatments Evaluated
Major Findings
Batt et al.
(2006)
Amherst, NY
Pharmaceuticals: iopromide; trimethoprim
Full-scale treatment plant:
Nitrification, SRT = 49 d
2 Laboratory-scale tests:
Nitrification was inhibited in one
bioreactor but not the other.
Both compounds more effectively
removed in nitrifying activated sludge
(SRT 49 d) than conventional
activated sludge (SRT 6 d).
Braghetta et al.
(2002)
Northern
California
17 PPCPs
1 allergy medicine: diphenhydramine
3 analgesics: acetaminophen; diclofenac;
ketoprofen
2 antibiotics: sulfamethoxazole;
trimethoprim
1 antiepileptic: carbamezipine
1 antihypertensive: diltiazem
1 blood glucose regulator: metformin
1 bronchodilator: albuterol
2 lipid regulators: fenofibrate; gemfibrozil
3 stimulant/stimulant metabolites: caffeine;
cotinine; paraxanthine;
2 ulcer medicine/histamine receptor
antagonists: cimetidine; ranitidine
• 2 WWTPs: Both use anthracite
filtration as tertiary treatment.
Plant A - conventional activated
sludge
Plant B - trickling filters/biofilters.
Note: SRTs were not specified.
Calculated removals using reverse
osmosis (RO) >90% for most
investigated PPCPs.
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Table 11-2. Study Design Parameters
Authors
Clara et al.
(2005a)
Clara et al.
(2005b)
Joss et al. (2004)
Study Location
Not specified
MBRand WWTP
(co-located) in
rural community
south-east of
Austria; 2 WWTPs
in Austria.
1 WWTP in
Kloten/Opfikon,
Switzerland; 1 in
Altenrhein,
Switzerland
Study Contaminants1
• EDC: bisphenol-A
• 3 Natural Estrogens: 17(3-estradiole (E2);
estrone (Ej); estriole (E3)
• 5 Pharmaceuticals: bezafibrate;
carbamazepine; diclofenac; 17a-
ethynylestradiole (EE2); ibuprofen
• 9 EDCs: bisphenol-A; nonylphenol;
nonylphenol diethoxylate; nonylphenol
monoethoxylate; nonylphenoxyacetic acid;
nonylphenoxyethoxyacetic acid; octylphenol;
octylphenol diethoxylate; octylphenol
monoethoxylate
• 2 Musk Fragrances: galaxolide; tonalide
• 8 Pharmaceuticals: bezafibrate;
carbamazepine; diazepam; diclofenac;
ibuprofen; iopromide; roxithromycin;
sulfamethoazole
• 2 Natural Estrogens: El, E2
• 1 Pharmaceutical: 17a-ethynylestradiole
(EE-,1
VCC2/
Treatments Evaluated
• 4 Full-scale plants: Most with N and
P removal, SRT10°cfrom 2 - >100 d.
• 1 Pilot plant: Membrane bioreactor
(MBR); ultrafiltration, SRT10°cfrom
22 - 88d.
• 4 Laboratory-scale tests:
1 SBR with SRT10°C of Id.
3 Bioreactors, N removal, SRT10°C
from 10-68 d.
• 3 Full-scale plants:
1 N and P removal, SRT 52-237 d, 6.8-
22.1 2C
1 C removal, SRT 2 d, 13.5 2 c.
1 nutrient removal, SRT 46 d, 10.4 2
C.
• 1 Pilot plant: (co-located at 1
WWTP): MBR, ultrafiltration
membrane, nitrification,
denitrification, SRT from 10 - 55 d,
5. 5- 27.2 2 C.
• 2 Full-scale plants: Both
conventional activated-sludge
treatment plants and
denitrification/nitrification.
1 included P removal, SRT of 10-12 d,
run in parallel w/ pilot MBR
1 run in parallel with fix-bed reactor,
Major Findings
• Enhanced removal with
increasing SRT for most
investigated compounds.
• Critical SRT can be defined,
beyond which removal rates
do not improve.
• No significant differences in
removal performances
between conventional
activated sludge systems and
MBR when operated at similar
SRTs.
See findings under Clara et al.
(2004a)
Biological degradation >90% of
investigated compounds observed
with nitrification and denitrification
(SRT of 12-15 d).
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Table 11-2. Study Design Parameters
Authors
Study Location
Study Contaminants1
Treatments Evaluated
Major Findings
SRTof22-24d.
1 Pilot plant: MBR, anaerobic &
anoxic compartments, aerobic
filtration compartments, 3
membrane filters (pore size, 0.4 |am,
0.1 urn, 0.04|am),SRTof30d.
Batch experiments: Aerated with
air for aerobic experiments and with
N for anaerobic experiments.
Kreuzingeret al.
(2004)
Austria
EDC: bisphenol-A
2 Musk Fragrances: galaxolide; tonalide
3 Natural Estrogens: El, E2, E,
9 Pharmaceuticals: bezafibrate;
carbamazepine; diazepam; diclofenac;
ethynylestradiol; ibuprofen; iopromide;
roxithromycin; sulfathometoxazol
4 Full-scale plants: 3 are single-
stage activated sludge, 1 is 2-stage;
all have P removal, 3 also include
nitrification/denitrification, SRT2o°c
of100d.
1 Pilot plant: Ultrafiltration, SRT20°C
of ll-41d.
Laboratory-scale plants: SRT20°c of
1- 26 d.
Enhanced removal with increasing
SRT for most investigated
compounds.
Khunjaret al.
(2007)
Virginia
EDC: EE,
Laboratory-scale plant: MBR
Note: Description provided in Yi (2006),
which was not reviewed.
Significant sorption onto
heterotrophic biomass versus
ammonia oxidizing biomass.
EE2 is transformed in the
presence of monoculture N.
europaea.
Inhibition of N. europaea by
EE2 is not expected at
concentrations typically found
in environment.
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Table 11-2. Study Design Parameters
Authors
Marttinen et al.
(2003)
Oppenheimer
(2007)
Study Location
Finland
Not specified
Study Contaminants1
Plasticizer: Bis(2-ethylhexyl) phthalate
20 PPCPs
• 2 antioxidants: butylated hydroxyanisol (BHA),
methyl paraben
• 1 detergent surfactant: octylphenol
• 2 fire retardants: tri(chloroethyl) phosphate;
tripheylphosphate
• 5 fragrances: ethyl 3-phenylpropionate;
galaxolide; methyl 3-phenylpropionate; musk
ketone; 3-phenylproprionate
• 2 germicides: triclosan; chloroxylenol
• 1 insect repellent: A/,A/-diethyl-3-
methylbenzamide (DEBT)
• 2 pharmacueticals: caffeine (includes non-
pharmaceutical); ibuprofen
• 1 plasticizer: butylbenzylphthalate
• 4 sunscreens: benzophenone; benzyl
salicylate; octyl methoxycinnamate;
oxybenzone
Treatments Evaluated
Full-scale plant:
Nitrification/denitrification, P removal, SRT
of 21 d at 18.52 C.
• 6 Full-scale plants: 1 activated
sludge, 1 high-purity oxygenated-
activated sludge, 4
nitrification/denitrification (1 of
which also includes granular
microfiltration/RO, and 3 include
UV), SRTsfrom0.5to30d.
• 2 Pilot-scale MBRs: 1 includes
nitrification/denitrification, SRTs of
14 Ki 1 ^ H
J-^T Ot J.J U .
Major Findings
• 94% removal from water
phase, primarily by sorption to
sludges.
• 29% removed during
nitrification/denitrification
(calculated)
• 32% during anaerobic
digestion (calculated)
• 32% remained in the digested
and dewatered sludge
(calculated).
• 62% overall removal rate
(calculated).
• Enhanced removal with
increasing SRT for most
investigated compounds.
• RO removed all 20
investigated PPCPs < detection
limit, including those not
consistently removed at SRTs
of 30 d.
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Table 11-2. Study Design Parameters
Authors
Study Location
Study Contaminants1
Treatments Evaluated
Major Findings
WERF (2005)
Not specified
EDCs (only categories were provided)
• Alkyl phenols (e.g., nonylphenol, bisphenol A)
• Metals/organometals (e.g., tributyl tin,
cadmium, lead)
• Organohalides
• PAHs/crude oil
• Plasticizers (e.g., phthalates)
• Steriods: naturally occurring; synthetic;
phytoestrogens
Indicated types of treatments only:
• Physical: primary settling, secondary
settling, filtration (e.g., sand).
• Biological: activated sludge, trickling
filter, anaerobic digestion.
• Chemical: chemical addition (e.g.,
phosphorus removal, polymers),
disinfection (chlorination, ozonation,
ultraviolet light).
• Advanced treatment: activated
carbon, membrane separation, RO,
and ion exchange.
Enhanced removal with
increasing SRT for most
investigated compounds.
Water temperature may
negatively impact the
effectiveness of EDC removal.
RO removal rates > 90% for
steroids, organohalides,
metals/organometals, and alkyl
phenols.
Yi and Harper
(2007)
Auburn, AL
EDC: EE,
Laboratory-scale: Nitrifying completely
mixed stirred tank reactor, SRT of 20 d.
Strong relationship between
nitrification and EE2 removal.
Contribution of heterotrophic
microorganisms to
biotransformation is unclear.
1 Although some of the study contaminants are listed as EDCs; others are suspected to be EDCs but require additional research (e.g., health effects studies) to make that
determination.
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