EPA/600/R-03/111
September 2003
Managing Urban Watershed
Pathogen Contamination
by
Joyce M. Perdek, Russell D. Arnone, and Mary K. Stinson
Water Supply and Water Resources Division
Urban Watershed Management Branch
Edison, New Jersey 08837
and
Mary Ellen Tuccillo
Oak Ridge Institute of Science and Education
Water Supply and Water Resources Division
Urban Watershed Management Branch
Edison, New Jersey 08837
National Risk Management Research Laboratory
Office of Research and Development
U.S. Environmental Protection Agency
Cincinnati, Ohio 45268
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Notice
The information in this report has been subjected to Agency peer and administrative
review and has been approved for publication as an EPA document. Mention of trade names or
commercial products does not constitute endorsement or recommendation for use.
11
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Abstract
This document is written as a resource for state and local watershed managers who have
the responsibility of managing pathogen contamination in urban watersheds. In addition it can
be an information source for members of the public interested in watershed mitigation efforts
aimed at reducing microbial contamination. It is written to support specific steps of the total
maximum daily load (TMDL) process for meeting water quality standards in urban watersheds.
The information provided can also support watershed evaluations conducted when disease
outbreaks occur in the absence of standards violations. The document discusses the regulation of
waterborne pathogens (Chapter 1), detection methods (Chapter 2), and combined sewer overflow
control technologies and stormwater best management practices (Chapter 3). The table below
identifies the steps of the TMDL process supported by each of the chapters.
The intent is to supplement the information included in the EPA document Protocol for
Developing Pathogen TMDLs, Office of Water, January 2001, EPA 841-R-00-002 guidance.
This document was developed using information collected through extensive literature reviews
by researchers in the Urban Watershed Management Branch (UWMB) of EPA's National Risk
Management Research Laboratory. The final document will be an official EPA report available
through the UWMB Internet site http://www.epa.gov/ednnrmrl/.
Steps of TMDL Process Supported by the Document Chapters
Steps of TMDL Process
TMDL Step 1:
Problem
Identification
TMDL Step 2:
Identification
of Water
Quality
Indicators and
Target Values
TMDL Step 3:
Source
Assessment
TMDL Step 4:
Linkage
Between
Water Quality
Targets
and Pollutant
Sources
TMDL Step 5:
Allocations
TMDL Step 6:
Follow-up
Monitoring
and
Evaluation
TMDL Step
7:
Assembling
the TMDL
Document Chapters
Chapter 1 . Pathogens of
Concern
Chapter 2. Detection Methods and
Alternate Indicator Organisms
Chapter 3.
Management
and Control of
Pathogens
111
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Foreword
The U.S. Environmental Protection Agency (EPA) is charged by Congress with
protecting the Nation's land, air, and water resources. Under a mandate of national
environmental laws, the Agency strives to formulate and implement actions leading to a
compatible balance between human activities and the ability of natural systems to support and
nurture life. To meet this mandate, EPA's research program is providing data and technical
support for solving environmental problems today and building a science knowledge base
necessary to manage our ecological resources wisely, understand how pollutants affect our
health, and prevent or reduce environmental risks in the future.
The National Risk Management Research Laboratory (NRMRL) is the Agency's center
for investigation of technological and management approaches for preventing and reducing risks
from pollution that threaten human health and the environment. The focus of the Laboratory's
research program is on methods and their cost-effectiveness for prevention and control of
pollution to air, land, water, and subsurface resources; protection of water quality in public water
systems; remediation of contaminated sites, sediments and ground water; prevention and control
of indoor air pollution; and restoration of ecosystems. NRMRL collaborates with both public
and private sector partners to foster technologies that reduce the cost of compliance and to
anticipate emerging problems. NRMRL's research provides solutions to environmental problems
by: developing and promoting technologies that protect and improve the environment; advancing
scientific and engineering information to support regulatory and policy decisions; and providing
the technical support and information transfer to ensure implementation of environmental
regulations and strategies at the national, state, and community levels.
This publication has been produced as part of the Laboratory's strategic long-term
research plan. It is published and made available by EPA's Office of Research and Development
to assist the user community and to link researchers with their clients.
Hugh W. McKinnon, Director
National Risk Management Research Laboratory
IV
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Contents
Notice ii
Abstract iii
Foreword v
List of Tables ix
List of Figures x
Acknowledgments xi
Chapter One Regulating Waterborne Pathogens 1-1
1.1 Introduction 1-1
1.2 Health Effects 1-4
1.2.1 Waterborne Disease Outbreaks 1-4
1.2.2 Pathogenic Bacteria of Concern 1-9
1.2.2.1 Campylobacter 1-9
1.2.2.2 E. Coli O157:H7 1-10
1.2.2.3 Legionellapneumophilia 1-11
1.2.2.4 Leptospira 1-11
1.2.2.5 Salmonella 1-11
1.2.2.6 Shigella 1-12
1.2.2.7 Vibrio cholerae 1-12
1.2.2.8 Yersinia entercolitica 1-12
1.2.3 Pathogenic Protozoa of Concern 1-13
1.2.3.1 Cryptosporidium 1-13
1.2.3.2 Cyclospora 1-15
1.2.3.3 Giardia lamblia 1-15
1.2.3.4 Entamoebahistolytica 1-16
1.2.3.5 Naegleriafowleri 1-16
1.2.4 Pathogenic Viruses of Concern 1-16
1.2.4.1 Adenoviruses 1-18
1.2.4.2 Astroviruses 1-18
1.2.4.3 Caliciviruses 1-18
1.2.4.4 Enteroviruses 1-18
1.2.4.5 Hepatitis A and Hepatitis E 1-19
1.2.4.6 Reoviruses 1-19
1.2.4.7 Rotaviruses 1-19
1.2.5 Pathogenic Helminth Worms 1-20
1.2.5.1 Nematodes 1-20
1.2.5.2 Cestodes 1-21
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Contents (cont.)
1.2.5.3 Trematodes 1-21
1.2.6 Pathogenic Fungi 1-22
1.3 Microbial Water Quality Standards 1-24
1.3.1 Clean Water Act 1-24
1.3.1.1 TMDL Description and Definition 1-25
1.3.1.2 Stormwater, Combined Sewer Overflow and Sanitary Sewer
Overflow Regulations 1-26
1.3.2 Safe Drinking Water Act 1-28
1.3.3 State Standards 1-29
1.3.4 Other Applicable Standards 1-29
1.3.4.1 Coastal Zone Act Reauthorization Amendments 1-29
1.3.4.2 Beaches Environmental Assessment, Closure, and Health
(BEACH) Program 1-31
1.4 Evaluation of Pathogen Indicators 1-31
1.4.1 Use of Indicators 1-32
1.4.2 Relationships between Indicators and Illness 1-34
1.5 Conclusions 1-37
References 1-39
Chapter Two Detection Methods and Alternate Indicator Organisms 2-1
2.1 Introduction 2-1
2.2 Detection Methods 2-2
2.2.1 Bacteria 2-2
2.2.1.1 Cultural and Enzyme-Based Methods 2-2
2.2.1.2 Immunological Methods 2-4
2.2.1.3 Genetic Methods (Gene Probes and PCR) 2-5
2.2.2 Viruses 2-6
2.2.2.1 Sample Concentration 2-6
2.2.2.2 Cultural Assay 2-8
2.2.2.3 Immunological Techniques 2-9
2.2.2.4 Gene Probes 2-9
2.2.2.5 PCR-based Methods 2-9
2.2.3 Cryptosporidium and Giardia 2-10
2.2.3.1 Immunofluorescence 2-10
2.2.3.2 Gene Probes and PCR-Based Methods 2-12
2.3 Alternative Indicator Organisms 2-12
VI
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Contents (cont.)
2.3.1 Clostridium perfringens 2-12
2.3.2 Bacteriophages 2-13
2.4 Microbial Source Tracking 2-13
2.4.1 Antibiotic Resistance Analysis 2-13
2.4.2 Molecular Methods 2-14
2.5 Conclusions 2-15
References 2-17
Chapter Three Management and Control of Pathogens 3-1
3.1 Introduction 3-1
3.2 Disinfection Technologies for Control of Pathogens 3-4
3.2.1 Introduction 3-4
3.2.2 WWF Disinfection Effectiveness 3-5
3.2.3 Requirement for a High-Rate Disinfection Process 3-6
3.2.4 Requirement for Suspended Solids Removal 3-6
3.2.5 WWF Disinfection Technologies 3-7
3.2.5.1 Chlorination and Dechlorination 3-7
3.2.5.2 Ultraviolet Light Irradiation 3-8
3.2.5.3 Chlorine Dioxide 3-10
3.2.5.4 Ozonation 3-12
3.2.6 Description of Disinfection Studies and Implementation Examples . . . 3-13
3.2.6.1 Disinfection Pilot Study at the 26th Ward WWTP Testing
Facility in New York City 3-13
3.2.6.2 Continuous Deflection Separation, Fuzzy Filter and UV
Treatment of SSO-Type Wastewaters: Pilot Study Results . . . 3-17
3.2.6.3 Advanced Demonstration Facility (ADF) in Columbus, GA . 3-20
3.2.6.4 Washington, DC. Northeast Boundary Swirl
Facility (NEBSF) 3-22
3.2.6.5 Birmingham, AL. UV Disinfection at Peak Flow WWTP . . 3-22
3.2.6.6 Oakland County, MI. Chlorine Disinfection at Acacia Park . 3-22
3.2.6.7 Bremerton, WA. UV Disinfection at CSO Treatment Facility 3-23
3.2.6.8 Disinfection of Collected Stormwater and Dry Weather
Urban Runoff 3-23
3.3 Best Management Practices (BMPs) for Control of Pathogens in Urban
Stormwater 3-24
3.3.1 Introduction 3-24
3.3.2 Structural BMPs 3-24
3.3.2.1 Ponds and Wetlands 3-28
vn
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Contents (cont.)
3.3.2.2 Sand Filters 3-30
3.3.2.3 Illicit Discharge Detection and Elimination 3-30
3.3.3 Nonstructural BMPs 3-31
3.3.3.1 Managing Waste from Resident Canada Geese 3-33
3.3.4 Effects of BMPs on Receiving-Water Quality 3-34
3.4 Conclusions 3-35
References 3-37
Vlll
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Tables
Table 1-1. U.S. Microbial Water Quality Assessments Summary - 1999 and 2000 1-2
Table 1-2. Outbreaks Associated with U.S. Natural Recreational Waters, 1986-2000 1-6
Table 1-3. Water Treatment Effectiveness on Pathogens 1-7
Table 1-4. Outbreaks Associated with Drinking Water from U.S. Surface
Sources, 1986-2000 1-7
Table 1-5. Waterborne Bacteria of Concern to Human Health and
Their Associated Diseases 1-10
Table 1-6. Waterborne Protozoans of Concern to Human Health and
Their Associated Diseases 1-14
Table 1-7. Waterborne Viruses of Concern to Human Health and
Their Associated Diseases 1-17
Table 1-8. Waterborne Helminths of Concern to Human Health and
Their Associated Diseases 1-21
Table 1-9. Waterborne Fungi of Concern to Human Health and
Their Associated Diseases 1-23
Table 1-10. Primary Contact Recreational Water Quality Criteria
for Microorganisms 1-33
Table 1-11. Key Points of Epidemiological Studies 1-35
Table 2-1. Summary of Detection Methods for Bacteria 2-7
Table 2-2. Summary of Detection Methods for Viruses 2-8
Table 2-3. Summary of Detection Methods for Cryptosporidium and Giardia 2-11
Table 3-1. Distinction between a Treatment Technology and a BMP for
Pathogen Control 3-3
Table 3-2. Cost Projection of Disinfection to be Implemented at the
Spring Creek Facility 3-18
Table 3-3. Stormwater BMP Effectiveness Data 3-26
Table 3-4. Results of Wetlands Effectiveness Studies on Secondary Sewage Effluent
at Pima County, AZ Constructed Ecosystem Research Facility 3-30
IX
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Figures
Figure 1-1. Microbial Pathogens Attributed to Cases of Illness from Exposure
to U.S. Surface Drinking Water Sources, 1986-2000 1-8
Figure 1-2. Microbial Pathogens Attributed to Outbreaks of Illness from Exposure
to U.S. Surface Drinking Water Sources, 1986-2000 1-8
Figure 3-1. Fecal Coliform % Removal Efficiency by BMP Type 3-28
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Acknowledgments
This report is a compilation of literature, reviewed, analyzed, and submitted by the U.S.
EPA's National Risk Management Research Laboratory, Water Supply and Water Resources
Division, Urban Watershed Management Branch (UWMB), Edison, NJ. This report contributes
to EPA's long term goal to provide the tools to restore and protect aquatic systems and to
forecast the ecological, economic, and human health outcomes of alternative solutions. The
report preparation commenced in September 2002, and the report was completed in September
2003. Annual performance requirements are fulfilled upon report completion.
Editor for this literature review is Joyce Perdek. Reviewers, who provided instrumental
comments, include Cecil Lue-Hing - President of Environmental and Water Resources Institute,
Sydney Munger and Charles Wisdom of Parametrix, Inc. (Kirkland, WA), Peter Swenson of U.S.
EPA Region 5, and Stephen Schaub and Don Waye of U.S. EPA Office of Water. In addition
several members of the UWMB provided valuable review, advice, and input. These reviewers
are Michael Borst, Carolyn Esposito, Chi-Yuan Fan, Richard Field, Bethany Madge, Thomas
O'Connor, Ariamalar Selvakumar, Daniel Sullivan, and Anthony Tafuri. Judy Norinsky of the
Environmental Careers Organization served as technical editor. The authors also wish to thank
fellow EPA researchers Mark Meckes, Frank Schaefer, and Joyce Simpson who offered their
microbiological expertise to assist in preparation of this and other UWMB pathogen projects.
XI
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Chapter One
Regulating Waterborne Pathogens
1.1 Introduction
Pathogens, disease causing microorganisms, are a major concern for managers of water
resources. Once in a water body, pathogens infect humans through contaminated fish and
shellfish, skin contact, or ingestion of water. Protection from pathogen contamination is most
important for waters designated for (1) recreation, (2) public water supplies, (3) aquifer
protection, and (4) protection and propagation offish, shellfish, and wildlife. These uses are
rigorously dealt with in Section 303(c) of the Clean Water Act (CWA) (U.S. EPA, 200 la). Data
on U.S. water bodies in violation of microbiological ambient water quality standards, established
by the states, for the years 1999 and 2000 are presented in Table 1-1.
The Maximum Contaminant Level Goals (MCLGs) established under the Safe Drinking
Water Act are zero for all pathogens. These goals conform to the position of the World Health
Organization (WHO) (1993):
"....there is no tolerable lower limit for pathogens, and water intended for consumption,
for preparing food and drink, or for personal hygiene should thus contain no agents
pathogenic for humans."
The WHO estimates that 13 million people die from waterborne infections each year. The
majority of these deaths occur in developing countries. However, in the U.S. approximately
900,000 cases of illnesses and 900 deaths occur each year as a result of microbial contamination
of drinking water (Warrington, 200la).
A pathogen may be a bacterium, protozoan, virus, worm, or fungi. Generally, waterborne
pathogens are in human and animal feces, and are deposited directly into water bodies or
transported to water bodies by overland flow and/or subsurface water flow. Urban pathogens are
transported by stormwater runoff, combined sewer and sanitary sewer overflows, and wastewater
treatment plant effluents. Pathogenic microorganisms originate from many animal species in
watersheds including wildlife, pets and companion animals, and agricultural animals. There is
increasing interest in the potential for molecular fingerprinting methods, also known as microbial
source tracking techniques, for identification of pathogen sources (Simpson et a/., 2002). The
majority of large scale pathogenic waterborne outbreaks in the past have been attributed to
human contamination or inadequacies at water treatment plants. The most current waterborne
1-1
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Table 1-1. U.S. Microbial Water Quality Assessments Summary - 1999 and 2000
Rivers and Streams
19% of U.S. river and stream miles assessed
39% of assessed river and stream miles impaired
Pathogens (bacteria) are leading cause of impairment
Agriculture is the primary source of impairment
Ocean Shorelines
6% of U.S. ocean shoreline miles assessed
14% of assessed shoreline miles impaired
Pathogens (bacteria) are leading cause of impairment
Urban runoff/storm sewers are primary source of impairment
Great Lakes Shorelines
92% of U.S. Great Lakes shoreline miles assessed
78% of assessed shoreline miles impaired
Pathogens (bacteria) are third leading cause of impairment
Contaminated sediments are the primary source of impairment
Estuaries
36% of U.S. estuarine square miles assessed
51% of assessed estuaries square miles impaired
Pathogens (bacteria) are fourth leading cause of impairment
Municipal point sources are primary source of impairment
Lakes, Reservoirs, and Ponds
43% of U.S. lake, pond and reservoir acres assessed
45% of assessed lake acres impaired
Pathogen (bacteria) are not a leading cause of impairment
Agriculture is the primary source of impairment
U.S. EPA, 2002a
outbreaks upon contact with contaminated recreational water bodies are attributed to human
fecal contamination or sewage (Levy etal., 1998; Upton, 1999).
Rosen (2000) identified the following characteristics of waterborne pathogens of
concern:
1. The organisms are shed into the environment in high numbers, or they are highly
infectious to humans at low doses.
2. The organism can survive and remain infectious in the environment for long periods or
they are highly resistant to water treatment.
3. Some types of bacterial pathogens can multiply outside of a host under favorable
environmental conditions.
Identifying the microorganisms causing water quality standard violations or waterborne
disease outbreaks is the first step in managing watershed microbial contamination. Emerging
1-2
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pathogens are disease agents that were unknown or not associated with water 10 to 20 years ago.
Many emerging pathogens are not new, but are only now associated with waterborne disease.
These novel disease-bearing microbes are engendered by a complex mixture including social,
political, economic, ecological, and technological factors, and are prone to arise among an
immuno-compromised population. Cryptosporidiumparvum, Legionella, and E. coll O157:H7
are preeminent waterborne emerging pathogens (Cliver, 2000; U.S. EPA, 2001b).
For U.S. water bodies not meeting state-established water quality standards for microbial
contaminants, a Total Maximum Daily Load (TMDL) must be developed. A TMDL is defined
as the maximum amount of a pollutant that a water body can receive and still meet the water
quality standard, and an allocation of that amount to the pollutant's sources. Usually, the TMDL
target level will be the numeric water quality criteria maximum for the microorganism for which
the standard was exceeded. In some cases, when the water quality standard does not sufficiently
reflect the use impairment, it is appropriate to develop and meet an alternative standard.
Examples of use impairments include waterborne disease outbreaks, degraded fisheries, and
restrictions on using the water body for the desired use of primary contact recreation. For these
situations, U.S. EPA recommends using a supplemental microorganism to provide additional
means for measuring attainment of designated or existing uses (U.S. EPA, 2001a).
This chapter provides information to support the first two steps of the seven step TMDL
process below. The information is also useful to support investigations of waterborne disease
outbreaks and management of water bodies not subject to the TMDL process.
TMDL Process
1. Problem Identification
2. Identification of Water Quality Indicators and Target Values
3. Source Assessment
4. Linkage Between Water Quality Targets and Pollutant Sources
5. Allocations
6. Follow-up Monitoring and Evaluation
7. Assembling the TMDL
The problem identification step's objective (U.S. EPA, 200la) is to:
Identify background information and establish a strategy for specific
303(d) listed waters that will guide the overall TMDL development
process. Summarize the pathogen-related impairment(s), geographic
setting and scale, pollutant sources of concern, and other information
needed to guide the overall TMDL development process and provide a
preliminary assessment of the complexity of the TMDL (what approaches
are justified and where resources should be focused).
The identification of water quality indicators and target values objective (US EPA, 200 la)
is to:
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Identify numeric or measurable indicators and target values that can be
used to evaluate the TMDL and the restoration of water quality in the
listed waterbody.
The information in this chapter on applicable numeric water quality standards, alternative
standards to support designated use, and evaluation of indicator microorganisms as water quality
criteria should be understood when undertaking problem identification in the TMDL process.
Information on pathogens causing waterborne disease outbreaks is provided as background and
may be most useful in situations where outbreaks occur.
1.2 Health Effects
This section discusses waterborne disease outbreaks and known waterborne pathogens.
The link between wet weather flow and outbreaks, and the data on pathogen related outbreaks
reported in the U.S. is presented. The following are detailed descriptions of bacteria, protozoa,
viruses, helminth worms, and fungi.
1.2.1 Waterborne Disease Outbreaks
Discharges of stormwater runoff, combined sewer overflows (CSOs), and sanitary
sewer overflows (SSOs) (all known as wet weather flows) to receiving waters create the
potential for disease outbreaks. Through climate and epidemiological records, Rose et al. (2000)
demonstrated a potential correlation between extreme precipitation events (the highest 20
percent of total intensity over a 20-year period) and waterborne disease outbreaks. The authors
found that statistically significant relationships could be identified between these precipitation
events and waterborne disease outbreaks due to contact with water from both surface and ground
water sources, although the relationship was much stronger for surface water outbreaks.
Swimming in contaminated marine and fresh recreational waters may result in a broad
spectrum of illnesses. Water bodies may be contaminated and continuously re-contaminated,
particularly if heavily used by people. For most pathogens warmer waters are more of a risk and
are pathogen reservoirs. Lack of flow and water stagnation allows pathogens to accumulate.
Swimming-associated disease outbreaks in natural U.S. waters between 1986 and 2000 due to
microorganisms are listed in Table 1-2. Exposure pathways of pathogens in recreational waters
are dermal contact, ingestion and inhalation resulting in skin, ear, eye, gastrointestinal, and
respiratory illnesses. Few studies other than those related to outbreaks have been conducted to
determine the etiological agents related to swimming associated illnesses (WHO, 1999). One
large-scale epidemiological study of swimmers in marine waters receiving stormwater runoff
involved interviewing over 15,000 individuals (Haile et al., 1999). Researchers reported higher
risks of upper respiratory and gastrointestinal infections for swimmers who swam (1) near storm-
drain outfalls, (2) in waters with high levels of single bacterial indicators and a low ratio of total
to fecal coliforms, and (3) in waters where enteric (intestinal) viruses were detected. These
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positive associations with adverse health effects indicate an increased risk of illness associated
with swimming in ocean water subject to untreated urban stormwater runoff. More than 1% of
the swimmers who swam in front of the outfalls were affected by fevers, chills, ear discharges,
vomiting, and coughing. Some studies attempting to link health effects to pathogen sources
yield inconclusive results. For example, seventeen E. coll O157:H7 cases led Perez Guzzi et al.
(2000) to investigate potential contamination from CSOs on California's Mar del Plata beaches.
Their investigation detected no E. coli O157:H7, although other strains of E. coll were detected
in 75% of the samples. None of the 98 strains detected in the outfalls were the strains that were
known to cause human illness.
Pathogens present in a watershed can enter the drinking water supply through stormwater
runoff, combined and sanitary sewer overflows, and illicit sanitary wastewater cross connections
into storm drains. Exposure pathways for pathogens in drinking water include ingestion, dermal
contact, and inhalation. Failures in water treatment systems, including the inability of
disinfection procedures to inactivate all pathogens, allow these microorganisms to remain in
finished water. Table 1-3 summarizes the effectiveness of water treatment processes on
waterborne pathogens. Giardia and Cryptosporidium caused the largest number of drinking
water-associated cases and outbreaks reported to the Center for Disease Control (CDC) from
1986-2000 (Table 1-4 and Figures 1-1,1-2). Although the drinking water treatment system met
state turbidity effluent requirements at all times immediately prior to and during the Milwaukee
Cryptosporidium outbreak in 1993, an assessment of the problem by a U.S. EPA investigative
team identified a potential link between high turbidity levels in the influent and the occurrence of
Cryptosporidium (Fox and Lytle, 1996). The American Society for Microbiology (ASM) reports
that outbreaks are associated with pathogen contamination of municipal water systems that
operate according to government standards, like Milwaukee. This indicates current
methodologies are unable to fully detect treatment system failures and water quality that will
adversely affect public health (Warrington, 200la).
Pathogen survival in aquatic environments affects their ability to cause illness. Many
environmental stressors effect survival, most notably sunlight intensity. Intense ultraviolet
sunlight over surface waters enhances bacterial die-off, therefore limiting serious bacterial
impacts (Chamberlin et. al., 1978). Bacteria in turbid waters and bottom sediments are not as
susceptible to sunlight as surface water microorganisms, and therefore survive longer. Protozoa
and viruses survive UV radiation better than bacteria (Johnson et. al., 1997). Pathogen survival
is also dependent on water temperature. Increased water temperature decreases the survival of
bacteria in surface water. Reduced cell metabolism in cold water enhances bacteria survival
(Terzieva etal, 1991). Protozoa and viral survival is also increased in cold water (LeChevallier
et al., 1991; Wait et al., 2000). Salinity (Johnson et al., 1997), competition and predation
(Rozen et al, 2001), and nutrient supply (Gauthier et al., 1989) are additional environmental
factors influencing die-off. Microbial survival is dependent on a combination of the above
factors. U.S. EPA compiled die-off rates of microbial indicators and pathogens in Table 6-1 of
Protocol for Developing Pathogen TMDLs (U.S. EPA, 200 la).
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Table 1-2. Outbreaks Associated with U.S. Natural Recreational Waters
1986-2000
Etiological Agent
AGr
Shigella spp.
Naegleria fovtferi
E.coliO157:H7
Schistosoma spp.
Cryptosporidium parvum
Norwalk-like
Giardia lamblia
Leptospira
E.coliO121:H19
unknown
Adenovirus 3
TOTAL
Cases#
1744
1618
16
336
203
649
257
83
389
11
4
595
5905
% of Cases
29.53
27.40
0.27
5.69
3.44
10.99
4.35
1.41
6.59
0.19
0.07
10.08
100
Outbreaks*
22
20
16
12
7
4
4
4
3
1
1
1
95
% of Outbreaks
23.16
21.05
16.84
12.63
7.37
4.21
4.21
4.21
3.16
1.05
1.05
1.05
100
# A case is defined as a disease occurrence from an etiological agent.
* An outbreak is defined as 1) greater than or equal to 2 persons experiencing a similar illness
after contacting the recreational water and 2) epidemiologic evidence that implicates the water
as the probable source of the illness.
** Acute gastrointestinal illness of unknown etiology.
Barwicketal., 2000; CDC and U.S. EPA, 1993; Herwaldtetal., 1992; Kramer etal., 1996; Lee
etal., 2002; Levine etal., 1990; and Levy etal., 1998.
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Table 1-3. Water Treatment Effectiveness on Pathogens
Pathogen Type
Bacteria
Protozoa
Viruses
Helminths
Fungus
Water Treatment and Effectiveness
Normal disinfection procedures using chlorine are
sufficient to kill bacteria
Multi-barrier approach including conventional
physical processes of sedimentation, coagulation
and filtration can remove 99% or better of most
protozoa. Chemical disinfection effectiveness is
minimal.
Conventional physicochemical processes of
sedimentation, coagulation, filtration and
chlorination effectively removes better than
99.99% of enteric viruses. The exception is the
Norwalk Virus which is resistant to chlorine
disinfection and relies on physical processes.
Conventional physicochemical processes of
sedimentation, coagulation, filtration and
chlorination effectively eliminate helminths.
Sub-micron filtration removes fungi. Fungi are
immune to normal levels of water chlorination but
are inactivated by UV or destroyed by ozone.
AWWA, 1999.
Table 1-4. Outbreaks Associated with Drinking Waterfrom U.S. Surface Sources
1986-2000
Etiological Agent
Campylobacter
Cryptosporidium parvum
Cyanobacteria-like
Giardia lamblia
Shigella sonnei
Ca. Jejuni
E.coli O157:H7
SRSV
AGI**
TOTAL
Cases#
250
419130
21
3424
1800
102
38
148
12169
437082
% of Cases
0.06
95.89
0.00
0.78
0.41
0.02
0.01
0.03
2.78
100
Outbreaks*
1
5
1
20
1
1
3
1
15
48
% of Outbreaks
2.08
10.42
2.08
41.67
2.08
2.08
6.25
2.08
31.25
100
# A case is defined as a disease occurrence from an etiological agent.
* An outbreak is defined as 1) greater than or equal to 2 persons experiencing a similar illness
** Acute gastrointestinal illness of unknown etiology.
Barwicketal.,2000; CDC and U.S. EPA, 1993; Herwaldt etal., 1992; Kramer etal., 1996; Lee
etal.,2002; Levineetal., 1990; and Levy etal., 1998.
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Figure 1-1. Microbial Pathogens Attributed to Cases of
Illness from Exposure to U.S. Surface Drinking Water
Sources, 1986-2000. Total Number of Cases =437,082.
nCryptosporidium parvum (95.9%)
nAcute Gastrointestinal II1
QGiardia lamblia (0.8%)
• Other (0.5%)
Illness of Unknown Etiology (AGI) (2.8%)
Figure 1-2. Microbial Pathogens Attributed to Outbreaks of Illness
from Exposure to U.S. Surface Drinking Water Sources, 1986-2000.
Total 48 Outbreaks.
QGiardia lamblia (41.7%)
D AGI (31.2%)
• Cryptosporidium parvum (10.4%)
QE. coliO157:H7(6.2%)
nCa. Jejuni (2.1%)
• Campylobacter (2.1%)
nCyanobacteria-like (2.1%)
nShigella sonnei (2.1%)
• Small Round Structured Virus
(SRSV)(2.1%)
Barwickef a/., 2000; CDC and U.S. EPA, 1993; Herwaldt et a/., 1992; Kramer et a/., 1996; Lee
et a/., 2002;Levine et a/., 1990; and Levy et a/., 1998
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1.2.2 Pathogenic Bacteria of Concern
Bacteria are unicellular microorganisms that exist as either free living organisms or as
parasites. Bacteria play a fundamental role in the decomposition and stabilization of organic
matter in nature and in biological sewage treatment processes. Bacteria range in size from 0.4 to
14 micrometers or microns (|im) in length and 0.2 to 1.2 jim in width. Many types of enteric
pathogenic bacteria occur in water supplies and in wastewater. The U.S. EPA (2000a; 2002a)
assessed bacteria as one of the leading causes of impairments to surface waters. With increasing
demands on water resources, the potential for contamination of surface and groundwater by
pathogenic enteric bacteria is expected to rise resulting in an increase in waterborne disease
outbreaks. Gastrointestinal illness, i.e., diarrhea, nausea, and cramps, is a common symptom of
infections caused by enteric waterborne bacteria. Some pathogens spread through the body from
the intestinal mucosa and cause systemic infections known as enteric fevers. One example of
this is typhoid fever. Chlorine disinfection is highly effective for most bacteria (AWWA, 1999).
Enteric bacteria tend to die off faster than strains indigenous to surface and groundwaters
because they are unable to compete successfully with natural microflora for low nutrient
concentrations (Sinclair and Alexander, 1984). However, some bacteria are able to adapt to low
nutrient concentrations by transforming to a viable but nonculturable (VBNC) state (Wang and
Doyle, 1998; Huq and Colwell, 1996). VBNC bacteria maintain metabolic activity and
infectiousness, but do not grow and multiply on culture plates, making them difficult to detect
with conventional methods. Enteric pathogenic bacteria transmitted by water and wastewater
include Campyiobacter, E.coli O157:H7, Leptospira, Salmonella, Shigella, Vibrio cholerae, and
Yersinia entercolitica. Legionellapneumophilia, while not enteric, is a pathogenic bacteria
distributed in the aquatic environment. Waterborne pathogenic bacteria of concern and their
associated diseases are presented in Table 1-5.
1.2.2.1 Campylobacter
Campylobacters of concern to the water industry are the "thermophilic" group. They
cause a variety of diseases in humans, principally acute diarrhea preceded by flu-like illness.
Campylobacter enteritis is principally a zoonotic disease, communicated from lower animals to
man under natural conditions. These bacteria are harbored in the intestines of domestic and
wild animals, particularly birds. Indirect transmission by contaminated water and food is the
most common infection mode. Campylobacter bacteria are killed by cooking procedures.
Campylobacter is now recognized as the cause of a common enteric bacterial infection in the
U.S. Over 21 cases for each 100,000 persons in the U.S. population (approximately 57,000
cases) are diagnosed each year (MMWR, 1999). Campylobacters are not found in water in the
absence of E. coll (AWWA, 1999).
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Table 1-5. Waterborne Bacteria of Concern to Human Health and Their Associated Diseases
Bacteria
Campylobacter
Escherichia coli O1 57:H7
(enteropathogenic)
Legionella pneumophilia
Leptospira (1 50 spp.)
Salmonella typhi
Salmonella (~ 1 700 spp.)
Shigella (4 spp.)
Vibrio cholerae
Yersinia entercolitica
Source
Bird feces
Cattle feces
Aquatic
environments
Urine of dogs,
livestock, wild
animals
Domestic and
wild animal
feces
Domestic and
wild animal
feces
Human feces
Asymptomatic
human feces
Animal feces
Disease
Diarrhea
Gastroenteritis
Legionellosis
Leptospirosis
Typhoid fever
Salmonellosis
Shigellosis
Cholera
Yersinosis
Effects
Acute diarrhea
Vomiting, diarrhea
Acute respiratory illness
Jaundice, fever (Weil's disease)
High fever, diarrhea, ulceration
of small intestine
Diarrhea, dehydration
Bacillary dysentery
Extremely heavy diarrhea,
dehydration
Diarrhea
Metcalf and Eddy, 1991.
1.2.2.2 E. Coli O157:H7
E.coli O157:H7 is a pathogenic strain of Escherichia coli belonging to the group
enterohemorrhagic E. coli. Human infection causes severe diarrhea and abdominal cramps. In
young children (under five years old) and the elderly, complications leading to life threatening
kidney failure can result (U.S. EPA, 2002b). The reservoir of this pathogen is primarily cattle.
This specific strain is an emerging cause of waterborne and foodborne illness. Fecally
contaminated water has been linked to recreational and drinking water outbreaks. An estimated
73,000 cases of infection and 61 deaths occur in the U.S. annually (CDC, 2001a). E. coli
O157:H7 was the responsible agent in the Cabool, MO disease outbreak that killed four people,
hospitalized 32 and caused diarrhea and other problems in 243 people (Geldreich et a/., 1992). It
is believed that breaks in drinking water mains resulted in low water pressure that allowed
contamination from nearby SSOs to enter the drinking water system. In 1999 this pathogen was
also responsible for the disease outbreak at a Washington County, NY fair due to contaminated
drinking water. Of the 781 people identified with illnesses related to this outbreak, 127 cases of
E. coli O157:H7 were confirmed by culture (Safefood News, 2000).
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U.S. public water systems must notify homeowners if the water is unsafe. Private well
owners should have their well tested periodically. Typically the well is tested for total coliform.
If the test is positive, the water is then tested for E. coli. If the E. coli is positive, the water
should not be consumed for drinking. U.S. EPA does not believe it is necessary for an owner of
a private well to test specifically for E. coli O157:H7 under normal circumstances because the
test is expensive and many labs do not have the expertise to perform this test (AWWA, 1999).
1.2.2.3 Legionella pneumophilia
Legionella is ubiquitous in the environment. The disease, legionellosis, is a severe
respiratory illness characterized by pneumonia. It is found typically in surface waters at
concentrations of 104 - 10s per liter and is now recognized as part of the natural environment
(Fliermans et a/., 1981). It has also proliferated in artificial environments such as cooling
towers, evaporative condensers, whirlpools, and hot water tanks. These environments act as
amplifiers or disseminators of legionellapneumophilia. In the U.S., 17,000 to 23,000 cases a
year are estimated. The largest outbreak occurred in Philadelphia, PA in 1976, where 220 cases
and 34 deaths were reported, and the source is unknown. Most outbreaks since 1976 have been
linked with hospital water distribution systems (AWWA, 1999).
1.2.2.4 Leptospira
Leptospira are spiral shaped bacteria. The induced disease, Leptospirosis or "Weil's
disease," first described in 1886, produces fever, headache, chills, malaise, vomiting, and
occasionally meningitis. This bacteria is transmitted through the urine of dogs, livestock, and
wild animals, and can contaminate natural water bodies, which then serve as sources of the
infection. Dogs are the major source for human infections. A vaccine is available for dogs but
not for humans. Between 100 and 200 documented cases per year occur in the U.S. (CDC,
200la). In the summer of 1998, 110 athletes competing in a triathlon in Illinois were diagnosed
with leptospirosis, and 23 needed hospital care. The outbreak was traced to Lake Springfield
(MMID, 1999).
1.2.2.5 Salmonella
Salmonella is a group of over 1,700 types of bacteria. There are three distinguishable
forms of salmonellosis, including gastroenteritis, enteric fever, and septicemia (characterized by
chills, fever, anorexia or loss of appetite) in humans. Gastroenteritis is characterized by
diarrhea, fever and abdominal fever. Enteric fever caused by Salmonella typhi is prolonged,
lasting from 7 to 14 days. Salmonella septicemia is characterized by chills, fever, anorexia, and
viable bacteria circulating in the blood known as bacteremia. Domestic and wild animals, and
humans are possible sources of Salmonella. Waterborne outbreaks of salmonellosis are normally
classified as acute gastrointestinal illness of unknown etiology. These outbreaks in the U.S. are
associated with poor quality source water and inadequate treatment and/or contamination of
distribution systems. In the U.S., over 40,000 cases of salmonellosis are reported each year, with
the incidence being about 17 cases per 100,000 people (CDC, 2000; CDC, 200Ib). The largest
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known waterborne incidence of this disease occurred in 1965 in Riverside, CA and affected
18,000 people. The water supply was blamed, but the source of contamination was never
determined (AWWA, 1999).
1.2.2.6 Shigella
Shigella is a genus of bacteria that causes sudden and severe gastroenteritis in humans,
known as shigellosis. Infected humans are the only significant reservoir. Waterborne outbreaks
result from fecal contamination of nonchlorinated private and noncommunity water supplies.
Septic tank contamination of wells, or cross-connections between wastewater and potable water
lines are commonly implicated in drinking water outbreaks. Recreational exposure to fecally
contaminated swimming areas is also prevalent (AWWA, 1999). Approximately 25,000
confirmed cases of shigellosis from all sources are reported in the U.S. each year. However,
many cases go undiagnosed, and 450,000 cases are estimated annually (Baer etal., 1999).
1.2.2.7 Vibrio cholerae
Over 130 groups of Vibrio cholerae have been studied. This bacteria is responsible for
the illness cholera, which produces acute diarrhea, dehydration, vomiting, shock, and possibly
death. Cholera is typically spread by poor sanitation. The most important reservoirs are
asymptomatic human carriers and diseased people who shed this bacteria in their feces.
Sporadic cases occur when shellfish are harvested and eaten raw from fecally polluted waters.
The excellent sanitation facilities in the U.S. are responsible for the near eradication of epidemic
cholera here. Cholera was reported in South America from 1991 to 1995, where it grew to
epidemic levels (1,099,882 cases and 10,453 deaths). Since this outbreak, most cases of cholera
in the U.S. have occurred among persons traveling from cholera-affected areas (CDC, 1995; U.S.
FDA, 2003a). Vibrio cholerae can also be present naturally in the environment, and natural
waters can be a source of this bacteria. This presence in the environment has been demonstrated
in the U.S. and Australia, where toxigenic strains survived in aquatic environments for years in
the total absence of fecal contamination (AWWA, 1999). Of particular concern is their presence
in warm, shallow, Gulf Coast waters, where oysters, as filter-feeders, concentrate these Vibrio
spp. organisms in their tissues (Hopkins et a/., 1997).
1.2.2.8 Yersinia entercolitica
Yersinia entercolitica is a facultative anaerobe (lives under either aerobic or anaerobic
conditions). This bacteria causes gastroenteritis, usually in children under seven years old,
characterized by fever, diarrhea, abdominal cramps, and sometimes vomiting. It is mainly
recognized as a foodborne pathogen, but may be found in sewage and polluted waters, and can
enter drinking water via pollution from these sources. Essentially, it is found where one might
encounter coliform organisms. However, Yersinia entercolitica is able to survive for longer
periods of time in aquatic environments (survival has been shown to grow at low temperatures
and survive for 18 months at 4°C) than fecal coliform. Therefore, this organism can be present
when the coliform indicator organisms are not (AWWA, 1997).
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1.2.3 Pathogenic Protozoa of Concern
Protozoa are one-celled animals varying in size from 2 to 100 jim. They live in many
animals and survive in cysts (protective shells) when outside of an organism. Protozoa
reproduce rapidly inside a host organism; therefore, ingestion of only a few by a human causes
disease. Once in water, protozoa can survive for several weeks, even longer if frozen in ice. The
waterborne pathogenic protozoans of greatest concern in countries with temperate climates are
Cryptosporidium and Giardia. Oocysts of Cryptosporidium and cysts of Giardia occur in
surface water, where their concentration is related to the level of fecal pollution or human waste
present. Oocysts and cysts are both very persistent in water and are very resistant to
disinfectants commonly used in drinking water treatment. In industrialized countries, outbreaks
of cryptosporidiosis and giardiasis are due to oocysts and cysts entering the drinking water
because of treatment failure, contamination of the source water, and/or leakage into the
distribution system (WHO, 1993).
Recently there is a growing concern regarding Cyclospora, especially in
nonindustrialized countries. Entamoeba histolytica and Naegleria fowleri are additional water-
transmitted intestinal parasites of concern worldwide due to their serious consequences. Table 1-
6 lists waterborne pathogenic protozoa of concern and their associated diseases.
1.2.3.1 Cryptosporidium
Cryptosporidium induces the disease cryptosporidiosis, which is capable of producing
unpleasant gastric and diarrheal illness (Rose, 1997). The parasite's transmittable stage is a 4 to
6 |im diameter spherical shaped oocyst which contains a hardy thick wall. The oocyst is spread
through the feces of infected humans and animals, including mammals, birds, reptiles, and fish.
Cryptosporidium is frequently waterborne in nature and infections have occurred through contact
with contaminated drinking water supplies, as well as zoonosis (animal person contact),
contaminated food, contaminated swimming pools, and other recreational waters. Oocysts may
be present in animal slurry spread on farmland as fertilizer. Consequently, runoff from rain
carries oocysts into streams, lakes, and other reservoirs. Sewage is another source. The
infective dose varies from less than 30 oocysts to as many as one million oocysts. There are six
species of Cryptosporidium, but only one species, Cryptosporidium parvum, found in animals, is
known to infect humans. Both known Cryptosporidium parvum genotypes can cause infections
in human beings. Genotype 1 has (so far) been found almost exclusively in humans, and is more
virulent than Genotype 2, which is found in a wide variety of animals, including humans (Xiao et
a/., 2001).
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Table 1-6. Waterborne Protozoans of Concern to Human Health and Their Associated Diseases
Protozoan
Cryptosporidium
Cyclospora
Entamoeba histolytica
Giardia lamblia
Naegleria fowleri
Source
Human,
animal, and
bird feces
Human feces
Human feces
Human,
animal, and
bird feces
Bird and
aquatic
mammal feces
Disease
Cryptosporidiosis
Cyclosporiasis
Amebiasis
(amoebic
dysentery)
Giardiasis
Meningoence-
phalitis (RAM)
Effects
Diarrhea, death in susceptible
populations
Diarrhea
Prolonged diarrhea with bleeding,
abscesses of the liver and small
intestine
Mild to severe diarrhea, nausea,
indigestion
Inflammation of brain and meninges
Fout, 2002; Metcalf and Eddy, 1991
States etal. (1997) found Cryptosporidium in treated sewage and CSO from an area
incorporating dairy farms. In an investigation of CSO in urban areas, Arnone etal. (2003)
reported essentially no Cryptosporidium in the two cities and three outfalls investigated. The
largest recorded outbreak of Cryptosporidiosis occurred in Milwaukee in 1993, where an
estimated 403,000 people were infected, and approximately 50 to 100 area residents with
compromised immune systems died prematurely (Blair, 1994; Hoxie et al., 1996). Another
significant Cryptosporidiosis outbreak occurred in Las Vegas in 1994, and infected 78 people,
most of whom had human immunodeficiency virus (HIV) infections (Roefer et al., 1996). At
present nothing other than the body's defense system can treat Cryptosporidiosis.
Cryptosporidium, therefore, poses some alarming public health problems, particularly for people
with weakened immune systems, especially acquired immunodeficiency syndrome (AIDS)
patients. These patients are prone to severe and protracted diarrhea which can persist for months
with considerable weight loss and mortality (Gerba etal., 1996; Rose, 1997).
A well-operated drinking water plant can physically remove only 99% of oocysts from
infected raw waters. Traditional processes such as coagulation, clarification, and filtration
remain the best defense against this parasite entering the water supplies. Encystment can protect
protozoa from drinking water disinfection efforts (Frey et al., 1998). U.S. EPA regulations
addressing this contaminant in drinking water supplies are discussed in U.S. EPA (200Ic) and
Chapter 3.
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1.2.3.2 Cyclospora
Cyclospora, species Cyclospora cayetanensis, are 8 to 10 jim in size. Disease symptoms
mimic those caused by cryptosporidiosis, including mild nausea, anorexia, abdominal cramping,
and diarrhea. Humans are the only natural host. Noninfectious Cyclospora oocysts are passed in
the feces of infected individuals. The unsporulated oocysts are usually transmitted via water and
require 7 to 15 days to sporulate and become infectious. Consumption of untreated water has led
to infection. During the spring of 1996 approximately 850 cases of cyclosporiasis were
confirmed in the U.S. and Canada. The infection lasts up to seven weeks. Symptoms typically
mimic those of cryptosporidiosis (AWWA, 1999).
1.2.3.3 Giardia lamblia
Giardia lamblia, also known as Giardia duodenalis and Giardia intestinalis, is the most
common cause of protozoa infection in humans. Sometimes referred to as "beaver fever,"
"hiker's disease," or "camper's disease," Giardia infection, or giardiasis, causes abdominal
cramps, diarrhea, and bloating. Giardia is found in humans, dogs, cats, pigs, sheep, beavers, and
many other domestic animals, as well as birds. Humans are usually infected by one particular
species of the many that exist, Giardia lamblia, which also causes infections in domestic and
wild animals. There are six strains of Giardia lamblia. The strain type is not consistently
associated with disease severity. Different individuals show various degrees of symptoms when
infected with the same strain (U.S. FDA, 2003b). The infection is transmitted by tiny spores or
egg-like cells called cysts measuring 9 to 12 jim in length. Watershed runoff and untreated and
treated sewage transport Giardia to lakes, rivers and other receiving water bodies. There is an
increase in Giardia infections during and after heavy rainfalls. Due to its thick wall, the Giardia
cyst can survive weeks or months in fresh water, although it is less hardy than the
Cryptosporidium oocyst (Rosen, 2000).
There have been over 20 outbreaks of waterborne Giardia in the U.S. from recreational
and surface drinking water contact between 1986 and 2000 (Barwick etal., 2000; CDC and U.S.
EPA, 1993; Herwaldt etal, 1992; Kramer etal, 1996; Lee etal, 2002; Levine et al., 1990; and
Levy et al., 1998). The infective dose for Giardia cysts may be between 10 and one million
viable cysts depending on the immune system of the host. Giardiasis can be treated with drugs,
including metronidazola, furazolidone, trinidazole, and paromomycin. Therefore, giardiasis is
not regarded as a fatal disease. Giardia infection occurs due to its reproduction in the digestive
system and attachment to the small intestine. After ingestion, the cyst passes through the
stomach to the duodenum where it hatches and produces two trophozoites, feeding configuration
of the parasite. The trophozoites measure 12 to 18 |im in length and adhere to the surface of the
mucous membranes of the small intestine. The trophozoites damage the membrane and inhibit
adsorption of nutrients that cause the disease giardiasis. The trophozoites then form cysts as they
pass along the small intestine and eventually pass out with the feces (Rosen, 2000). Many
individuals are asymptomatically affected by Giardia, as demonstrated by a CDC study of a
population who consumed water heavily contaminated with Giardia due to malfunction in the
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water disinfection system. Only 11 percent of the exposed population developed symptoms even
though 46 percent had the organism in their stools (Rockwell, 2002).
Giardia cysts are removed fairly readily by conventional drinking water treatment
processes, such as coagulation, settlement, and rapid filtration. A well operated treatment plant
utilizing coagulation, clarification, and filtration should remove 99.9% of Giardia from the
water. Disinfection with chlorine is ineffective due to the cyst's thick wall. Current research
indicates that irradiation with ultraviolet (UV) light is the most promising form of Giardia
disinfection or inactivation (U.S. EPA, 2001c).
1.2.3.4 Entamoeba histolytica
Entamoeba histolytica causes the disease known as amoebiasis, characterized by
dysentary, chronic colitis, and liver abscess. Infected humans, particularly asymptomatic
carriers, are the only reservoirs of significance. Waterborne outbreaks in the U.S. are rarely
documented. The most dramatic outbreak in the U.S. was the 1933 Chicago World's Fair
outbreak caused by contaminated drinking water, infecting 1,400 individuals and causing 58
deaths (Warrington, 200Ib). An estimated 40 million people worldwide develop this disease
annually, and the mortality is estimated at 40,000.
1.2.3.5 Naegleria fowleri
Naegleriafowleri causes an acute rapid occurring disease of the central nervous system
primary amebic meningoencephalitis (PAM). This disease is characterized by severe headache,
fever, and coma leading to death within 3 to 10 days after the onset of symptoms. Birds and
aquatic mammals such as beavers, otters, and muskrats are reservoirs for this pathogen.
Naegleriafowleri is found free in the environment, specifically soils, freshwater, and sewage. It
enters the body through the nasal passage and travels along the nerves to the meninges. It
comprises both nonpathogenic and pathogenic strains (Geldreich, 1996).
1.2.4 Pathogenic Viruses of Concern
Viruses are a group of infectious agents that require a host to survive. They use the host
cell's reproductive mechanism to replicate. After replication, and subsequent death of the host
cell, viral particles are spread to neighboring cells. This results in infection to the individual.
Viruses are the smallest and most basic life form, ranging in size from 0.02 to 0.09 jim. The
virus protein or lipoprotein covering enables it to survive for long periods and adhere to surfaces
(AWWA, 1999). Table 1-7 lists the viruses of concern to human health via exposure to water
and their associated diseases (Fout, 2002; Metcalf and Eddy, 1991).
The viruses most significantly affecting water quality and human health are enteric
viruses which are found in the gastrointestinal tract of infected individuals. These viruses are
excreted in the feces of infected people and may directly or indirectly contaminate water
intended
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Table 1-7. Waterborne Viruses of Concern to Human Health and Their Associated Diseases
Virus
Adenovirus (48 serotypes; types
40 and 41 are of primary
concern)
Astrovi ruses
Calicivirus (e.g., Norwalk,
Norwalk-like and Sapporo,
Sapporo-like viruses)2
Enterovirus (66 types, e.g.,
polio, echo, encephalitis,
conjunctivitis, and Coxsackie
viruses)
Hepatitis A
Hepatitis E1
Reovirus
Rotavirus
Source
Humans
Humans
Humans
Humans
Humans
Humans,
pigs
Humans
Humans
Disease
Respiratory
disease,
gastroenteritis
Gastroenteritis
Gastroenteritis
Gastroenteritis,
heart anomalies,
meningitis
Infectious
hepatitis
Infectious
hepatitis
Gastroenteritis
Gastroenteritis
Effects
Acute respiratory disease,
pneumonia, conjunctivitis,
gastroenteritis
Vomiting, diarrhea
Vomiting, diarrhea
Respiratory illness, polio,
common cold
Jaundice, fever
Jaundice, fever
Vomiting, diarrhea
Vomiting, diarrhea
1 Hepatitis E is an emerging virus that has caused large outbreaks of
U.S.
2 Norovirus and Sapovirus are the new genus names for the Norwalk-
Fout, 2002; Metcalf and Eddy, 1991
infectious hepatitis outside the
•like and Sapporo-like viruses.
for drinking. Enteric viruses multiply only within living cells. They take over a living cell and
use the cell's reproductive mechanism to replicate. Most waterborne virus disease outbreaks in
the U.S. are caused by sewage contamination of untreated or inadequately treated private and
semipublic water supplies. Conventional physicochemical water treatment processes of
coagulation-flocculation and filtration remove up to 99% of most enteric viruses. Disinfection of
water with free chlorine, chlorine dioxide, ozone, and UV light radiation can achieve 99.9%
enteric virus inactivation. Norwalk virus is the exception; this virus is very resistant to chlorine
and other disinfection measures (AWWA, 1999).
The predominant enteric viruses of concern are enteroviruses, rotaviruses, hepatitis A and
E, caliciviruses, adenoviruses, reoviruses, and astroviruses. Each consists of subgroups totaling
more than 140 different enteric viruses known to cause numerous illnesses that include diarrhea,
fever, hepatitis, paralysis, meningitis, and heart disease. Some viral infections are asymptomatic
(AWWA, 1999).
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1.2.4.1 Adenoviruses
Human adenoviruses may cause acute respiratory disease, pneumonia, epidemic
conjunctivitis, and acute gastroenteritis in children. Human adenovirus is not pathogenic to
animals, and animal adenovirus is only pathogenic to the species of origin. Hurst et al. (1989)
found that adenoviruses are transmitted through recreational and drinking water. Jiang et al.
(2001) found adenoviruses at beach locations in southern California, with concentrations ranging
from 880 to 7,500 plaque-forming units (PFU) per liter of water.
1.2.4.2 Astroviruses
Astroviruses have a unique star-shaped surface when viewed by a negative-stain electron
microscopy. These viruses produce symptoms similar to those caused by rotaviruses, including
vomiting, diarrhea, and mild dehydration. There have been no reports on infectivity for animals
and there are no known reservoirs for these viruses. Astroviruses are primarily transmitted by
the fecal-oral route (AWWA, 1999). Immunity to astrovirus infection is not well understood.
Young children and the institutionalized elderly are usually the populations that develop
symptomatic infection. This suggests that the antibody is acquired late in childhood and
provides protection through adult life until the elder years when this protection is diminished.
1.2.4.3 Caliciviruses
Norwalk and Norwalk-like viruses are caliciviruses, also known as small round-
structured viruses (SRSV). Norwalk virus, the prototype SRSV was first isolated in 1972 at an
elementary school in Norwalk, Ohio. The genus name for the Norwalk-like virus is now called
Norovirus. Another calicivirus is the Sapporo-like virus, of which the genus name is now
Sapovirus. These viruses produce vomiting in children and diarrhea in adults. In the U.S., 40%
of the outbreaks of gastroenteritis in adults are attributed to these two viruses. Humans are the
only reservoir for caliciviruses. Norwalk and Norwalk-like viruses are transmitted by ingestion
of fecally contaminated material. Infections have been associated with ingestion of surface
water contaminated by fecal material, ingestion of groundwater contaminated by septic drainage,
and swimming in sewage-contaminated waters. Outbreaks also occur following consumption of
shellfish harvested from waters contaminated with human sewage. Oysters, clams, and other
shellfish filter virus particles from contaminated water and accumulate them in their tissues
(AWWA, 1999).
1.2.4.4 Enteroviruses
Enteroviruses cause a wide variety of illnesses, ranging from polio to the common cold.
Non-polio enteroviruses are second only to the rhinoviruses, which cause the common cold, and
are the most common viral infectious agent affecting humans. Infected persons who become ill
develop respiratory flu-like symptoms. Less commonly, some people develop viral meningitis.
Humans are the only natural hosts for these viruses. Enterovirus infection is prevalent upon
exposure to human fecal contamination in a variety of sources, including groundwater, marine
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waters, shellfish, crops irrigated with sewage, and where spray irrigation of sewage is practiced.
The enteroviruses cause an estimated 10-15 million symptomatic infections each year in the U.S.
and many more asymptomatic infections (AWWA, 1999). Noble and Fuhrman (2001) found
enteroviruses in 32% of beach waters sampled from the Santa Monica Bay, CA suggesting the
potential for enterovirus infection through recreational contact.
1.2.4.5 Hepatitis A and Hepatitis E
Hepatitis A (HAV) causes the disease known as "infectious hepatitis," which is an acute
inflammation of the liver. Hepatitis E (HEV) also causes infectious hepatitis and is nearly
indistinguishable from HAV. Humans are the main reservoir for infectious hepatitis and shed
the virus in their feces. Pigs are also a reservoir for HEV. Direct and indirect person-to-person
contact are the primary HAV exposure mechanisms. Fecally contaminated drinking and bathing
water, and shellfish harvested from fecally contaminated waters serve as reservoirs and
transmission pathways. In the U.S., 20,000 to 30,000 cases of the HAV cases are reported
annually. HEV is rare in the U.S. although it is widespread in other parts of the world (AWWA,
1999). Although mortality from infections caused by these viruses is comparatively low, the
disease may be severe and incapacitating. Case fatality rates of 20-40% are on record for HEV
infections in pregnant women (Grabow, 1997). Vaccines are only available for HAV, and no
meaningful treatment is available for any of the hepatitis viruses, making disease control
dependent solely on preventing transmission. In Oocee, FL SSOs periodically flooded a mobile
home park during heavy rains and caused occasional outbreaks of hepatitis A from November
1988 to April 1989; 39 cases were identified among residents and 100 cases were linked to food
handlers living in the park. The initial reports by public health officials attributed the outbreaks
to poor personal hygiene rather than to the SSOs. It took four years for officials to determine the
connection between the SSOs and the outbreaks (Vonstille etal., 1993).
1.2.4.6 Reoviruses
Reovirus infections are mostly subclinical or very mild. These viruses are found in the
respiratory and enteric tracts. They lack a direct association with a specific human disease.
Reports associate reoviruses with a host of different diseases such as juvenile onset diabetes,
fever, rash, respiratory disease, pneumonia, eye infections, and meningitis. Reoviruses are
ubiquitous in nature and are commonly found in sewage and fecally polluted waters. Their main
source is human excreta.They are the most commonly isolated viruses from water and are easily
recognized (AWWA, 1999).
1.2.4.7 Rotaviruses
Rotavirus is responsible for 3.5 million cases of diarrhea and 125 deaths per year in the
U.S. Humans and animals are the primary reservoirs for rotaviruses. Infections occur mostly in
infants and children under two years old. Rotaviruses are predominately transmitted by the
fecal-oral route. Most rotavirus infections occur in the winter in temperate climate. Rotavirus is
responsible for 30 to 50 percent of U.S. hospitalizations for diarrhea in children under five years
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old. Immunocompromised patients and the elderly are also susceptible to this virus. This virus
is believed to be the cause of gastroenteric waterborne illnesses, and has been detected in
freshwater and sewage. Associated attacks have been documented. One such outbreak occurred
among users of the Vail, CO community water system where a very high adult attack rate of
43.8% was recorded (AWWA, 1999).
1.2.5 Pathogenic Helminth Worms
Helminth worms refer to those of the cestode (tapeworm), trematode (fluke), and
nematode (roundworm) groups. There are many waterborne helminth worms that are pathogens.
Although these worms are multicellular with complex reproduction systems and life cycles,
helminths are more completely understood than many other group of pathogens. Many of them
require invasion of a host which results in illness, damage, and sometimes the death of the host
to complete its life cycle. These are known as parasites. Some helminths normally live and
replicate in the natural environment or in other species. These infect humans when conditions
are right for causing disease, but do not complete their life cycle in the individual. They are not
parasites but opportunistic pathogens that can spread from individual to individual and from
individuals to water. Over one billion people worldwide are infected annually with intestinal
helminths. In the U.S., intestinal helminth disease has been largely eliminated due to improved
sanitation. Table 1-8 lists the waterborne pathogenic helminth worms of concern and their
associated diseases (Warrington, 200Ic). This information on helminths is for informational
purposes only since helminth infection is not prevalent in the U.S.
1.2.5.1 Nematodes
Nematodes (roundworms) are elongated, unsegmented, cylindrical worms, distinguished
by both sexes. Ascaris lumbricoides is the largest nematode, reaching 35-cm long in adult
females and 30-cm long adult males. These worms are the most common human helminth
infection, afflicting over 800 million people annually worldwide. The highest prevalence is in
tropical and subtropical regions, especially in areas with inadequate sanitation, but infections are
also reported in rural areas of the southeast U.S. Ascaris lumbricoides causes parasitism in the
human intestine, known as ascariasis. Ancylostoma duodenale and Necator americanus are
hookworms that infect over 350 million humans worldwide each year. Iron deficient anemia
accompanied by cardiac complications are the most common symptoms of this disease known as
ancylostomiasis. Trichuris trichiura is the third most common roundworm found in humans.
Distribution is worldwide, with infections more frequent in tropical areas and environments with
poor sanitation systems. The disease known as trichuriasis is most frequently asymptomatic.
Heavy infections, especially in children, can cause gastrointestinal problems such as abdominal
pain, diarrhea, rectal prolapse and possible growth retardation (Warrington, 200Ic).
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Table 1-8. Waterborne Helminth of Concern to Human Health and Their Associated Diseases
Helminth
Ascaris lumbricoides
Ancylostoma duodenale
Necator americanus
Trichuris trichiura
Taenia solium
Diphyllobothrium latum
Schistosoma haematobium
Schistosoma intercalation
Schistosoma japonicum
Trichobilhartzia spp.
Source
Human feces
Human Feces
Human feces
Human feces
Pigs
Fish
Snails, human
feces
Snails, human
feces
Snails, human
feces
Snails,
waterfowl,
aquatic animals
Disease
Ascariasis
Ancylostomiasis
hookworm
Trichuriasis
Taeniasis
Diphyllobothriasis
Schistosomiasis
Schistosomiasis
Schistosomiasis
Swimmer's itch
Effect
Asymptomatic, respiratory
problems
Anemia
Anemia
Gastrointestinal problems
Intestinal disturbance
Anemia, diarrhea
Diarrhea, lesions, cystitis
Diarrhea, lesions, cystitis
Diarrhea, lesions, cystitis
Open sores and lesions
in skin
Warrington, 2001 c; WHO, 1999
1.2.5.2 Cestodes
Cestodes (tapeworms) are flat segmented worms that are hermaphroditic (having both
male and female reproductive organs). These parasites are found in the gut and acquired by
ingesting contaminated food and water. Intermediate hosts (cattle, pig, fish) ingest these
waterborne parasites. Infections are passed to humans who eat the meat of the intermediate
hosts. The head or scolex attaches to the wall of the gut, and segments up to 25 meters long
called proglottids are attached behind the head. The proglottids are full of eggs and as new ones
are produced, the old one, containing up to 1,000,000 eggs, detach and are shed with the feces.
Annually Diphyllobothrium latum (fish tapeworm) infects 10,000,000 people, and Taenia solium
(pork tapeworm) infects 6,500,000 people worldwide (Warrington, 2001c).
1.2.5.3 Trematodes
Trematodes (flukes) have complex life cycles, usually involving a snail and some other
intermediate host such as fish, crustaceans and sheep. Flukes are unsegmented, flat, leaf-shaped
worms having a variety of organ systems. Most flukes are hermaphroditic. They attach to the
host by means of an oral sucker and a ventral sucker. Flukes, as adults, infect either the portal
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blood vessels, intestines, liver, or lungs of humans. Schistosoma haematobium, Schistosoma
japonicum, and Schistosoma intercalatum penetrate the skin of humans and cause
schistosomiasis. This disease worldwide affects approximately 195,000,000 people annually and
is responsible for 15,000 deaths. Swimmer's itch is a skin rash caused by parasites of birds and
mammals. In the U.S., the species is normally Trichobilharzia spp. and is reported in areas
along the migratory bird flyways where avian hosts are common. Waterfowl, mainly ducks and
geese, are the hosts of the schistosomes that cause schistosomiasis (Warrington, 200 Ic).
1.2.6 Pathogenic Fungi
Fungi, including yeasts and filamentous species or molds, are ubiquitously distributed
heterotrophic (requiring complex organic compounds for metabolic synthesis) organisms found
in lakes, ponds, streams, estuaries, marine environments, wastewaters, rural and urban
stormwater runoff, and aquatic sediments. Normal healthy individuals rarely suffer from
waterborne fungal diseases; it is the immunocompromised individuals that are at risk of fatal
fungal infections. Fungi are not generally problems in drinking water. They may present
problems when water is used for bathing and recreational activities (Warrington, 200Id).
Fungi are aerobic, multicellular, nonphotosynthetic organisms having organized nuclei,
usually rigid walls, and lack chlorophyll. The presence of fungi in stream water represents soil
runoff because nearly all zoopathogenic fungi exist saprobically (feeding on dead or decaying
material) with soil as their natural reservoir. Fungi pathogenic to humans are found in pools and
beaches and in accompanying washing facilities. Table 1-9 lists the waterborne pathogenic fungi
of concern and their associated diseases. Along with bacteria, fungi are the main organisms
responsible for the decomposition of carbon on earth. Without the presence of fungi to break
down organic matter, the carbon cycle (the process by which carbon is exchanged between
organisms and the environment) would cease to exist and organic matter would start to
accumulate. Aquatic species include fungi that are transiently present in water, terrestrial fungi
that disperse in water, and species that function entirely within water. Unpolluted stream water
has a large number of species representing true aquatic fungi (species possessing flagellated
zoospores and gametes), aquatic Hyphomycetes, and soil fungi. Moderately polluted waters may
carry cells of all three types, but with fewer true aquatic fungi and aquatic Hyphomycetes, and
soil fungi are more numerous. Heavily polluted water has large numbers of soil yeast-like fungi
(descend a/., 1998).
The association between fungal densities and organic loading implies that fungi may be
useful indicators of pollution. However, no single species of fungi has been identified as
important in this role. Fungi are found in potable water and on the inner surface of distribution
system pipes. They either survive water treatment or they enter the system after treatment.
Having survived treatment fungal spores can remain viable for extended periods. For instance,
pathogenic spores of Histoplasma capsulatium remain highly infective to mice after 400 days
(descend a/., 1998).
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Table 1-9. Waterborne Fungi of Concern to Human Health and Their Associated Diseases
Fungus
Aspergillus fumigatus
Candida albicans
Geotrichum candidum
Histoplasma capsulatum
Pseudallescheria boydii
Rhinocladiella mansonii
Source
Soil, decaying
organic matter
Raw
wastewater
Sewage, soil
Bird droppings
Sewage, soil
Soil, plants,
water
Disease
Pulmonary
aspergillosis
Candidiasis
Geotrichosis
Histoplasmosis
Eumycotic
mycetoma
Chro mo mycosis
Effects
Inflammation of bronchi and
lungs
Infection of moist cutaneous
areas of body
Infection of mouth, respiratory
tract
Respiratory infections
Infection of the cutaneous and
subcutaneous tissues
Skin lesions
Clescerief a/., 1998
Aspergillus fumigatus., an agent of an inflammatory and destructive disease of the
bronchi and lungs known as pulmonary aspergillosisi, has been found almost everywhere on
every conceivable type of substrate, especially soil and decaying organic debris. Candida
albicans, found in raw wastewater, wastewater treatment plant effluents, or contaminated water,
is a parasitic fungus that can infect the mouth, skin, intestines, and vagina. Geotrichum
candidum is found worldwide in sewage, soil, and water as well as in plants, cereals, and dairy
products. It is responsible for geotrichosis, an infection of the mouth, respiratory tract and
digestive tract
Histoplasma capsulatium is a thermally dimorphic fungus found in nature. Soil
contaminated with bird droppings is the common natural habitat for Histoplasma capsulatium.
The spectrum of the disease varies from an acute pulmonary infection to a chronic pulmonary
disease. Pseudallescheria boydii is a causal agent of the slow, destructive infection of cutaneous
and subcutaneous tissues recognized as eumycotic mycetomas. It is found in soil, sewage,
contaminated water, and the manure of farm animals. Rhinocladiella mansonii, also known as
Exophiala mansonii, is a saprophyte found in soil, plants, water, and decaying wood material. It
is responsible for the disease chromomycosis, or skin lesions. (Clesceri et al., 1998).
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1.3 Microbial Water Quality Standards
Section 303(d) of the CWA requires each state to develop a list of impaired waters.
Impairment is determined relative to state water quality standards for a given water body.
Section 303(d) requires states to develop pollutant-specific TMDLs for each impaired water
body. The objective of establishing and implementing the TMDL is to achieve the water quality
target for that pollutant class. Generally, the target established is the state's water quality
standard, which is based on U.S. EPA's recommended water quality criteria. There are
exceptions, however, as explained in the Protocol for Developing Pathogen TMDLs (U.S. EPA,
200 la):
"In some cases, the water body of concern has a numeric water quality standard that
might not appropriately or sufficiently reflect the use impairment, and the use of a
supplementary indicator or set of indicators might provide additional means for
measuring attainment of designated or existing uses."
Examples of use impairments include waterborne disease outbreaks, degraded fisheries, and
restrictions on using water body for the desired use of primary contact recreation. In these cases,
U.S. EPA (200la) recommends using a supplementary microbial indicator or pathogen that
reflects the problem affecting the designated use and establishing the target receiving-water
concentration and TMDL accordingly. The sections below describe the water quality legislation
relevant to selecting the water quality target.
1.3.1 Clean Water Act
Section 303(c) of the CWA requires states to adopt water quality standards that take into
account the designated uses for the water body. Standards must be set to support those
designated uses and be based on U.S. EPA's recommended water quality criteria developed
pursuant to Section 304(a) (U.S. EPA, 2000a).
Section 305(b) of the CWA requires all states and jurisdictions to:
1. assess the health of their waters and the extent to which their waters support water quality
standards
2. identify the pollutants and sources contributing to water quality impairments
3. analyze the economic and social costs and benefits of achieving the goals of the CWA
4. submit reports every two years to the U.S. EPA describing water conditions
The CWA Section 305(b) further requires U.S. EPA to summarize reports from the states and
other jurisdictions and convey this information to Congress biennially, currently the National
Water Quality Inventory 2000 Report (U.S. EPA, 2002a). The assessments reported under
Section 305(b) are used to identify and prioritize water quality problems within states. This
report is developed from impaired water bodies identified in accordance with Section 303(d).
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Section 303(d) of the CWA identifies waters that do not or are not expected to meet water
quality standards after implementation of water pollution controls.
Once the 303(d) list is prepared, states develop TMDLs. The U.S. EPA and state water
programs are currently working on sequencing water quality monitoring to determine
appropriate water quality standards to support the full range of water quality management (U.S.
EPA, 2000a). The sequence of activities consists of:
1. Characterizing waters for the 305(b) assessment
2. Using the subset of waters identified as not supporting water quality standards to develop
303(d) lists
3. Identifying source contributions
4. Developing TMDLs
5. Implementing source controls
6. Performing follow up monitoring to evaluate the effectiveness of source controls and to
track trends in water quality improvements
1.3.1.1 TMDL Description and Definition
TMDLs are developed for a variety of pollutants, such as: (1) oxygen depleting
substances, (2) nutrients, (3) sedimentation and siltation, (4) bacteria and pathogens, (5) toxic
organic chemicals and metals, (6) pH, (7) habitat and hydrologic modification, (8) suspended
solids, (9) noxious aquatic plants, (10) oil and grease, and (11) salinity and mineralization (U.S.
EPA, 2000a). A maximum pollutant amount, TMDL, is required for each water body that cannot
be improved by simply enforcing the minimum required source treatment. A TMDL sets a
pollution cap. The cap is a formula representing the maximum amount of a pollutant (pathogen
in the case of this document) that a water body can receive and still meet water quality standards.
The sum of the allowable contributing point and nonpoint sources must not exceed this cap. A
TMDL is the sum of the individual wasteload allocations for point sources and load allocations
for nonpoint sources and natural background with a margin of safety (CWA section
303(d)(l)(c)). The TMDL, expressed in terms of mass (or organism counts for microorganisms)
per time, can be described generically by the following equation:
TMDL = LC = X WLA + £ LA + MOS (U.S. EPA, 2001a)
where: LC = loading capacity, the greatest loading a water body can receive without
exceeding water quality standards
WLA = wasteload allocation, the portion of the TMDL allocated to existing or
future point sources
LA = load allocation, the portion of the TMDL allocated to existing or future
nonpoint sources and natural background
MOS = margin of safety which is provided implicitly through analytical
assumptions or explicitly by reserving a portion of loading capacity
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TMDLs are developed to meet applicable water quality standards. Standards may be
expressed by numeric water quality targets, narrative criteria for designated uses such as
drinking water, recreation, fish and wildlife habitat. The numeric target may be equivalent to a
numeric water quality standard as found in U.S. EP'A''s Bacterial Water Quality Standards Status
Report (U.S. EPA, 1998a), or it may represent a quantitative interpretation of a narrative
standard. U.S. EPA's water quality criteria provide guidance for the amount of pathogen
degradation a water body can accommodate while still supporting the specific uses.
Understanding when a water body is most vulnerable to pathogen contamination is
critical to developing load reduction scenarios that will result in attainment of water quality
standards. When an impairment is the result of contributions from sewage treatment plants and
industrial point sources, it is usually most pronounced at low flows. This is because point source
contributions are relatively constant over time. When stream flow is low, these point source
discharges constitute a relatively large proportion of the total stream flow. If an impairment is
more pronounced at higher flows, the pollutant is associated with wet weather, i.e., stormwater
runoff, combined sewer overflows, and some sanitary sewer overflows (U.S. EPA, 2000b).
As stated in the introduction, the seven components of the TMDL development process
are:
1. Problem identification
2. Identification of water quality indicators and targets
3. Source assessment
4. Linkage between water quality targets and sources
5. Allocations
6. Follow-up monitoring and evaluation
7. Assembling the TMDL
These seven components, discussed in detail in the document Protocol for Developing Pathogen
TMDLs (U.S. EPA, 200la), provide a guidance and framework for the TMDL development
process. TMDL calculations and allocations are a legally required components of the TMDL
package submittal.
1.3.1.2 Stormwater, Combined Sewer Overflow and Sanitary Sewer Overflow
Regulations
1.3.1.2.1 Stormwater
Stormwater runoff is generated from land and impervious areas during rainfall and snow
events. These runoffs often contain pollutants, including pathogens, that adversely affect water
quality. Polluted stormwater runoff is a leading cause of impairment to nearly 40 percent of
water bodies in the U.S. that do not meet water quality standards (U.S. EPA, 2002c).
Urbanization drastically alters the stormwater quality and quantity through hydraulic
modifications. These modifications include catchbasins, inlets, curb and gutter, gutter and
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downspouts, storm sewers, ditches, lined channels, culverts, and pavement. Storm water travel
time is reduced and flow velocity is increased as compared to the original natural conditions
(Field and Sullivan, 2003). Studies show a linear relationship when the runoff volume is
regressed against watershed imperviousness (Schueler, 1987).
Most stormwater discharges are considered point sources and require a National Pollutant
Discharge Elimination System (NPDES) permit. U.S. EPA developed Phase I of the NPDES
stormwater program in 1990 in response to the 1987 amendments to the Clean Water Act. Phase
I requires operators of medium and large municipal separate storm sewer systems (MS4s) to (1)
obtain a NPDES permit, (2) develop a stormwater management program to prevent pollutants
from being washed by stormwater into the MS4, then discharged from the MS4 into local water
bodies. A medium MS4 is a system that is located in an area with a population between 100,000
and 249,999. A large MS4 is a system that is located in an area with a population of 250,000 or
more. In addition, Phase I requires a NPDES permit for stormwater discharges from
construction areas that disturb five acres or greater of land. The Phase II Final Rule was signed
by the U.S. EPA Administrator on October 29, 1999. Phase II requires NPDES permit coverage
to (1) stormwater discharges from certain regulated small MS4 (communities less than 100,000,
primarily those located in urbanized areas), and (2) small construction areas disturbing between
1 and 5 acres of land. Best management practices (BMPs) are the primary method to control
stormwater discharges (U.S. EPA, 2002c). Use of BMPs for microbial contaminants is discussed
in Chapter 3.
1.3.1.2.2 Combined Sewer Overflow
Combined sewer systems (CSSs) convey sanitary wastewater and stormwater through a
single pipe to a publicity owned treatment works for treatment prior to discharge to surface
waters. The U.S. EPA 2001 Report to Congress (U.S. EPA, 2001d) reports that CSSs are found
in 32 states (including the District of Columbia). CSSs are concentrated in older communities in
the Northeast and Great Lakes regions. This report documents 772 CSO communities with a
total of 9,471 CSOs that are identified and regulated by 859 NPDES permits. Approximately 30
percent of the CSS communities have populations greater than 75,000, and approximately 30
percent have total service populations of less than 10,000. The annual CSO discharge is
estimated at 1,269 billion gallons per year. CSO receiving water are distributed to 43 percent
rivers, 38 percent streams, five percent oceans, estuaries and bays, and two percent other waters
(ditches, canals, unclassified waters). CSOs are a source of impairment for 12 percent of
assessed estuaries (in square miles) and two percent of assessed lakes (in shore miles).
Overflows occur during moderate or heavy rainfall when capacity is exceeded. CSOs deposit
water with varying concentrations of sanitary wastewater onto public areas, potentially resulting
in a range of adverse health effects (Colford et a/., 1999).
1.3.1.2.3 Sanitary Sewer Overflow
SSOs are discharges of raw sewage from municipal sanitary sewer systems. Most SSOs
are associated with wet weather conditions, when sanitary systems receive stormwater in-flow or
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infiltrating groundwater through cracks. The SSO may occur during extreme hydrologic events
in many separate sanitary systems, even though systems are intended to collect and contain all
the sewage that flows into them. U.S. EPA estimates 40,000 SSOs annually (U.S. EPA, 2001e).
Discharges to waters of the U.S. from municipal sanitary sewer systems are prohibited, unless
authorized by an NPDES permit. There are approximately 19,000 municipal sanitary sewer
collection systems in the U.S. The U.S. EPA proposed SSO Rule clarifies and expands
requirements for these collection systems, with the premise of reducing SSOs (U.S. EPA,
2002d).
1.3.2 Safe Drinking Water Act
In the U.S., both the 1986 and 1996 Amendments to the Safe Drinking Water Act
(SOWA) focused attention on source water protection and its role in protecting public water
supplies. Developed to support SDWA implementation, U.S. EPA's Surface Water Treatment
Rule (SWTR) (U.S. EPA, 1989), Interim Enhanced SWTR (IESWTR) (U.S. EPA, 1998b) and
Long Term 1 Enhanced Surface Water Treatment Rule (LT1ESWTR) (U.S. EPA, 2002e) are
designed to prevent waterborne diseases caused by viruses, bacteria, and the protozoans Giardia
lamblia and Cryptosporidium, which are present in varying concentrations in most surface
waters. These rules set unenforceable maximum contaminant level goals (MCLGs) of zero for
pathogens in treated drinking water because exposure to them at any level poses a health risk.
Rather than establishing a maximum contaminant level (MCL) for these contaminants in
drinking water, U.S. EPA opted instead to impose a treatment requirement. Utilities using
surface water must filter and disinfect the water to provide at least 99% removal/inactivation of
Cryptosporidium, 99.9% of Giardia, and 99.99% of viruses. Unfiltered public water systems
must have watershed control programs to reduce the sources and limit the migration of these
pathogens into raw waters. The U.S. EPA has established a MCL for total coliform detection at
no more than 5.0% of samples per month (for water systems that collect fewer than 40 routine
samples per month, no more than one sample can be total coliform-positive per month). Every
sample that has total coliform must be shown to contain no fecal coliform (U.S. EPA, 2003a).
In November 2001, the U.S. EPA issued the pre-proposal draft of the National Primary
Drinking Water Regulations: Long Term 2 Enhanced Surface Water Treatment Rule
(LT2ESWTR) (U.S. EPA, 2001f). The purposes of the LT2ESWTR are to improve control of
microbial pathogens (specifically Cryptosporidiuni) in drinking water and to address risk trade-
offs with disinfection byproducts. The LT2ESWTR provisions are:
1. source water monitoring for Cryptosporidium with reduced monitoring requirements for
small systems
2. additional Cryptosporidium treatment for filtered systems based on source water
Cryptosporidium concentrations
3. inactivation of Cryptosporidium by all unfiltered systems
4. disinfecting, profiling and bench marking to assure continual levels of microbial
protection while public water systems take the necessary steps to comply with new
disinfection byproduct standards
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5. covering, treating, or implementing a risk management plan for uncovered finished water
reservoirs
6. criteria for a number of treatment and management options to meet additional
Cryptosporidium treatment requirements
1.3.3 State Standards
The CWA allows states, tribes, and other jurisdictions to develop their own water quality
standards to protect their waters. At a minimum, they include the swimmable and fishable goals
of the CWA. States must submit their standards for U.S. EPA approval. Monitoring data are
compared to the standards for water quality assessment and decisions on whether to list waters as
impaired under the CWA Section 303(d). Water quality standards have three critical elements
(U.S. EPA, 2002a):
1. Standards should state designated uses that water quality should support, such as
recreation, aquatic life, fish consumption, drinking water supply, industry, agriculture,
and navigation. Each use has unique set of water quality criteria that must be met for the
use to be realized.
2. State water quality criteria are both numeric and narrative. Numeric criteria are
thresholds required to support a beneficial use. Narrative criteria describe conditions that
must be maintained to support a designated use.
3. States provide an antidegradation statement intended to prevent waters currently in
degraded condition from further deteriorating, and minimizing deterioration of high
quality water.
The U.S. EPA is actively promoting its goal of ensuring that all states and tribes update
their bathing beach standards. After state standards are set, states assess their waters to
determine the degree to which these standards are met. The U.S. EPA Bacterial Water Quality
Standards Status Report (U.S. EPA, 1998a) is an overview of the bacterial water quality
standards that have been adopted by states for their marine and fresh recreational waters. The
U.S. EPA Bacterial Water Quality Standards For Recreational Waters, Freshwater and Marine
Waters (U.S. EPA, 2003b) is an update of the 1998 status report. The 1998 and 2003 reports
indicate that many states have adopted E. coli and enterococci standards; however, many of these
states are still regulating according to fecal coliforms requirements while building databases for
E. coli and enterococci monitoring data.
1.3.4 Other Applicable Standards
1.3.4.1 Coastal Zone Act Reauthorization Amendments (CZARA)
In 1990, Congress passed legislation to protect America's coasts from runoff pollution. It
created the Coastal Nonpoint Program, also known as Section 6217 of the CZARA of 1990.
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Section 6217 requires the participating U.S. coastal states and territories to establish effective
programs to control and prevent polluted runoff into coastal waters. Section 6217 is the first
national program to tackle, in a comprehensive and enforceable fashion, the problem of coastal
nonpoint pollution. Currently, every eligible coastal state participates in the Coastal Zone
Program with the exception of Illinois (U.S. EPA, 2003c).
1.3.4.2 Beaches Environmental Assessment, Closure, and Health (BEACH)
Program
The U.S. EPA announced on May 23, 1997 the Beaches Environmental Assessment,
Closure, and Health (BEACH) Program (U.S. EPA, 2003d). The BEACH program goal is to
reduce the risk of infection to users of the nation's recreational waters. High levels of pathogens
in recreational waters can increase human exposure through inhalation, ingestion, and body
contact. Scientific studies document the presence of disease-carrying bacteria, viruses, and other
pathogens present in local beach water, primarily from sewage and stormwater. The BEACH
Program focuses on the following areas to meet the program goals of improving public health
and environmental protection programs and providing the public with information about the
quality of their beach water:
1. Strengthening the beach standards and testing
2. Providing faster laboratory test methods
3. Predicting pollution
4. Investing in health and methods research
5. Informing the public
Congress subsequently passed the BEACH Act in October 2000 that authorizes U.S. EPA
to award grants to eligible states, tribes, and territories to develop and implement beach water
quality monitoring programs at coastal and Great Lakes recreational waters near beaches. These
grants further support the development and implementation of programs to inform the public
about the risk of exposure to disease-causing microorganisms in the waters at the nation's
beaches. Nearly $10 million in grants were awarded on April 4, 2003 to 35 eligible states (U.S.
EPA, 2003d).
In addition to the BEACH Program initiatives, U.S. EPA is involved with the following
activities in other programs to make its waters cleaner and safer for swimming:
1. Assist communities to build and properly operate their sewage treatment plants
2. Work to end sewage overflows in communities with outdated sewer systems
3. Implement a national stormwater program to reduce urban runoff
4. Adapt the CZARA
5. Improve sewage disposal from recreational vessels
National implementation of strong, consistent beach programs will provide the public
with important information about the quality of their beach water and allow the public to make
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decisions on when and where to swim. The U.S. EPA operates a Website called "Beach Watch"
that provides information about the water quality at our nation's beaches, local protection
programs, and other beach related programs. The "Beach Watch" Website is updated as new
information becomes available, and is available at http://www.epa.gov/waterscience/beaches/
(U.S. EPA, 2003d).
1.4 Evaluation of Pathogen Indicators l
This section addresses the complexities associated with using indicators to evaluate water
quality for microbiological contamination. Measuring microbial indicators is less expensive,
easier, and more common than measuring pathogens directly. Generally, standards set are based
on the presence or concentrations of bacterial indicator organisms and the designated use of the
water body. In many cases, the sources for the increased indicator microorganism levels are
known and establishing the corresponding TMDL is appropriate. However, as stated earlier,
there are exceptions, which include cases where numeric water quality standards do not exist and
there is identifiable impairment of designated uses (U.S. EPA, 200la). A target value that
reflects attainment of the designated uses should be selected and the TMDL developed to meet it.
If there are existing standards that do not adequately address the designated use, a supplementary
indicator should be used along with the existing standard. The supplementary indicator would
provide additional means for measuring attainment of designated or existing uses.
Indicator organisms and monitoring programs are limited in their ability to predict
pathogen presence and health risks. Watershed managers need to understand the complexities
associated with using indicators in order to protect public health. Therefore, they may need to
conduct monitoring and management measures in addition to those required by the TMDL
process. For example, in the event that standards are not met and the pollution source is not
evident, watershed evaluations may be necessary to identify causes of elevated indicator
concentrations and pathogens present. Alternatively, evaluations would also be warranted when
outbreaks occur in the absence of standards violations. The sections below provide background
on the indicators used as water quality standards in the U.S.
1.4.1 Use of Indicators
Bacterial indicators were originally adopted to alert public health officials to the presence
of human fecal contamination in drinking water supplies. Because of the time, labor, expense,
complexity, and analytical limitations associated with directly analyzing for a variety of specific
pathogens, bacterial indicator use has remained the mainstay of microbial water quality
monitoring for decades. A good indicator organism is present when the pathogens of concern
1 Much of this section is excerpted from Monitoring Pathogens in the Watershed: Indicator
Organisms and Detection Methods, submitted for peer-reviewed publication by M.E. Tuccillo and J.M.
Perdek, and Investigating Watershed Microbial Pathogen Contamination to Manage Public Health Risks,
submitted for peer-reviewed publication by J.M. Perdek, M.E. Tuccillo, and S.M. Wankel.
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are present and is easy and inexpensive to detect. It must also occur in much greater numbers
than the pathogens and should be at least as resistant to adverse environmental conditions as the
pathogens (Armon and Kott, 1996). Indicators are used as the basis for water quality criteria
developed to support designated uses, such as primary contact recreation and drinking water
supply.
Current U.S. microbial water quality standards are based on criteria developed between
20 and 40 years ago. Based on the Federal Water Pollution Control Administration's (now U.S.
EPA) technical advisory report on water quality criteria (Federal Water Pollution Control
Administration, 1968), the U.S. EPA recommended that states adopt as a bathing water quality
standard fecal coliforms not to exceed 200 organisms/100 mL (U.S. EPA, 1976). There was
concern, however, that insufficient data existed to support this decision (National Research
Council, 1972). The U.S. EPA conducted research between 1972 and 1979 to reexamine the
question of health effects related to swimming in sanitary wastewater-polluted waters. Central to
this program, several epidemiological-microbiological studies concluded that fecal coliforms
density showed little correlation with swimmer gastrointestinal illness (Cabelli, 1983; Dufour,
1984). Dufour (1984) reported a high correlation between gastrointestinal illness in fresh waters
and both Enterococcus and E. coli concentrations, while Cabelli (1983) reported a high
correlation between Enterococcus concentrations and gastrointestinal illness in marine waters.
Based on this research, the U.S. EPA (1986) revised its recreational water quality criteria to the
indicators and concentrations shown in Table 1-10. Fresh water criteria are 33/100 mL for
Enterococcus or 126/100 mL for E. coli. The marine water criterion for Enterococcus is 35/100
mL. While most experts agree that E. coli and Enterococcus are superior indicators than fecal
coliform, fecal coliform is still widely used because of its historic use. The information in Table
1-10 also portrays differences in microbial recreational water quality standards in countries
around the world. These differences reflect not only the differences in indicator suitability for
various geographical areas, water types, and pathogen sources, but also the diversity of opinions
as to the most appropriate indicator.
Present approaches to regulating and monitoring recreational waters for pathogen
contamination in the U.S. and worldwide suffer from limitations. To examine the issue, experts
representing U.S. EPA and the WHO met in November 1998 in Annapolis, MD. The outcome of
this meeting, known as the "Annapolis Protocol," is an improved approach for regulating
recreational waters that better reflects health risk and yields an enhanced scope for effective
management intervention (WHO, 1999). The protocol consists of a classification scheme that
provides for assigning a level of risk to a beach area and indicates the management and
monitoring actions likely to be appropriate. Risk is determined using a combination of
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Table 1-10. Primary Contact Recreational Water Quality Criteria for Microorganisms
Country
U.S. Marine Water 1
U.S. Fresh Water 1
Australia
Brazil2, Peru3
Canada 4
Colombia, Ecuador
Cuba1
European Union 5
France 8
Hong Kong (marine) 3
India
Israel 9
Japan
Mexico
Poland
Uruguay
Yugoslavia
Enterococcus
/100ml_
35
33
—
—
350/L
—
—
90%<100
—
—
—
—
—
—
—
—
—
E. coli
/100ml_
—
126
—
—
2,000/L
—
—
—
—
180
—
—
—
—
<1000
—
—
Total Coliform
/100ml_
—
—
—
80%<5000
—
1000
1000
95%<1 0,000 6
80%<500 7
<2000
—
500
—
1000
80%<100010
100%<1 0,000
—
—
2000
Fecal Coliform
/100ml_
—
—
300
80%<1000
—
200
200, 90%<400
95%< 2,000 6
80%<1007
<500
—
—
1000 6 100 7
—
—
—
<50011,<100012
—
1 Geometric mean of at least 5 samples equally spaced over a 30-day period.
2 "Satisfactory" waters, samples obtained in each of the preceding five weeks.
3 Geometric mean of 5 most recent concentrations.
4 Geometric mean of at least 5 samples, taken during a period not to exceed 30 days. When
experience has shown that greater than 90 percent of fecal conforms are E. coli, either fecal
conforms or E. coli may be measured.
5 The European Union also has a Salmonella requirement of 0/L, a fecal streptococcus guideline of
100/100ml_, and an enterovirus requirement of 0/1 OL.
6 Mandatory 7 Guideline
8 France also has a fecal streptococci requirement of <1 00/1 00 ml.
9 Israel uses World Health Organization Guidelines. 10 At least 5 samples per month.
11 Geometric mean of at least 5 samples. 12 Not to be exceeded in at least 5 samples.
Corbett et a/., 1993; Council of European Communities, 2000; Fattal et a/., 1987; Federal-Provincial
Working Group, 1992; Govt. of India MOEF, 2000, Ho and Tarn, 1998; Salas, 1998 (in WHO 1999); and
U.S. EPA, 1986.
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microbiological indicator monitoring data and an inspection-based assessment of the
susceptibility of an area to direct influence from human fecal contamination.
1.4.2 Relationships between Indicators and Illness
The most common microbiological indicator of pathogens is the presence of coliform
bacteria, which are commonly found in the enteric tracts of humans and other warm-blooded
animals. Testing for these organisms is relatively fast, easy, and inexpensive. If these bacteria
are detected in sufficiently high concentrations, then there is a high probability of contamination
of human fecal matter, which may contain pathogens. Coliform bacteria are members of the
Enterobacteriaceae family and include species of Enterobacter, Klebsiella, Citrobacter, and
Escherichia. Fecal coliforms, a subset of total coliforms, are defined by their ability to grow at
elevated temperature (44.5°C). They are associated with the enteric tracts of warm-blooded
animals, whereas total coliforms can include bacteria from cold-blooded animals and soil
organisms. Fecal coliforms include the familiar Escherichia coli. Although most E. coli (and
most coliforms) are not harmful, some strains, including E. coli O157:H7, are pathogenic as
discussed in Section 1.2.2.2.
Studies evaluating the use of coliform bacteria as indicators of fecal contamination have
shown mixed results (Table 1-11). Some researchers have reported favorably on coliform
testing, often in conjunction with other microorganisms, as an indicator of pathogens.
Epidemiological research in the United Kingdom concluded that the European Union's
recreational water testing requirements, which include total coliforms, fecal coliforms, and other
organisms, adequately protect the health of swimmers in coastal waters (Pike, 1994). Several
other investigators (Corbett et al., 1993; Ferley et al., 1989; Haile et al., 1999; Seyfried et al.,
1985) have reported correlations between fecal coliforms concentrations and incidence of
general morbidity or total illness.
Numerous studies, however, have found that fecal coliforms or total coliforms
concentrations do not correlate well with illness (Calderon et al., 1991; Cheung etal., 1990;
Fattal etal., 1987; Kay etal, 1994; Kueh etal, 1995; andMcBride etal, 1998). As stated
above, the U.S. EPA epidemiological-microbiological studies conducted in the 1970s concluded
that fecal coliform densities showed little or no correlation with gastrointestinal illness among
swimmers (Cabelli, 1983; Dufour, 1984). Although fecal coliforms are primarily associated with
the enteric tracts of warm-blooded animals, Dufour (1984) suggests that many bacteria in the
environment fit the description of fecal coliforms but do not come from gastrointestinal sources.
Thus, these bacteria are of questionable use as fecal indicators.
The relationships between coliform bacteria and pathogenic microorganism densities are
also problematic. Studies have found poor or no correlations between coliform densities and
pathogenic bacteria or enteric viruses in environmental waters (Metcalf et al., 1995; Morinigo et
al., 1992; Olivieri etal., 1977). Griffin etal. (1999) found enteroviruses in canals of the Florida
Keys although none of the sites studied violated state water quality standards for total coliforms
or fecal coliforms. Coliform bacteria are poor indicators for the pathogenic protozoa
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Table 1-11. Key Points of Epidemiological Studies
Author
Cabelli, 1983
Fattalefa/.,
1987
McBride et a/.,
1998
Cheung et a/.,
1990
Kuehefa/., 1995
Kayefa/., 1994
Corbett et a/.,
1993
Haileef a/., 1999
Pike, 1994
Dufour, 1994
Ferley et a/.,
1989
Seyfried ef a/. ,
1985
Calderon ef a/.,
1991
Country
U.S.
Israel
New
Zealand
Hong Kong
Hong Kong
United
Kingdom
Australia
U.S.
Great
Britain
U.S.
France
Canada
U.S.
Water/
Discharge
Types
Marine/Sewage
Marine/Raw
Sewage
Marine/Treated
Marine/Sewage
and Stormwater
Marine/Sewage
and Stormwater
Marine/Various
Marine/Primary
treated
Marine/
Stormwater
Marine
Fresh/Sewage
Fresh/Untreated
Sewage
Fresh/Not Stated
Fresh/Animal
Nonpoint Source
Indicator Best
Correlated with
Swimming-Associated
Gastrointestinal Illness
Enterococcus; EC to a
lesser extent
Enterococcus
Enterococcus
EC (r=0.73) and
Enterococcus (r=0.60)
Turbidity, Clostridium
perfringens, Aeromonas
spp., Vibrio cholerae
(non-01)
FS
None found
FC, Enterococcus, EC,
Viruses, Distance from
Storm Drain
TC and enteroviruses;
FS in cohort study
EC and Enterococcus
FS
None found at 0.05 level
of significance; FS at
0.069 level of
significance
Staphylococci; bather
density
Indicator Best Correlated
with Other Swimming-
Associated Illnesses
N/A
N/A
N/A
N/A
none found
N/A
FC for cough, ear symptoms,
eye symptoms, fever
FC for skin and respiratory
symptoms, enterococcus for
skin symptoms, EC for eye,
ear and skin symptoms,
Viruses for fever, chills, eye,
and respiratory symptoms
none found
N/A
FC for general morbidity; FC,
Aeromonas, and
Pseudomonas aeruginosa for
skin diseases
Total staphylococcus
(strongest), FC and FS
(weakest) for total illness; total
staphylococcus for eye and
skin illnesses
N/A
FC - fecal coliform, FS - fecal streptococcus, EC - E. coli, TC - total coliform
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Cryptosporidium parvum and Giardia lamblia. According to Craun et al. (1997), coliforms have
been detected during most waterborne outbreaks caused by bacteria and viruses, but during
relatively few outbreaks caused by protozoa. Drinking water disease outbreaks, most notably
from Cryptosporidium and Giardia, have occurred in water systems that did not violate the U.S.
EPA-issued MCL for total coliforms (Craun et al., 1997). Rose et al. (1991) found no
associations between either Cryptosporidium or Giardia and the total coliform or fecal coliform
indicator in drinking water sources. Because coliform bacteria are killed more easily by
disinfection than viruses or protozoa, coliform absence in evaluating disinfection does not
guarantee the absence of health risk (Toranzos and McFeters, 1997).
Several epidemiological studies support the use of E. coli and the Enterococcus group as
indicators of fecal contamination. Enterococcus, a subgroup of fecal streptococci, is frequently
found in the human digestive tract. They are tolerant of a wide range of environmental
conditions and are easy to culture. These characteristics render the Enterococcus group a
promising indicator. As stated above, research in the U.S. has demonstrated high correlations
between Enterococcus densities and gastrointestinal illness among swimmers in fresh (Dufour,
1984) and marine (Cabelli, 1983) waters. Epidemiological research in Israel and New Zealand
demonstrated strong relationships between Enterococcus densities and the incidence of illness
among swimmers in marine waters receiving raw sewage (Fattal et al., 1987) and treated sewage
discharges (McBride et al., 1998). Other investigators have also reported positive relationships
between the incidence of gastrointestinal illness among marine water swimmers and
Enterococcus densities (Cheung et al., 1990; Haile et al., 1999). Dufour (1984) found a high
correlation between E. coli and illness in fresh waters.
For cases of nonhuman pollution sources, i.e., runoff from rural areas containing animal
waste, using E. coli or Enterococcus may not be the best approach. In a study of swimmers in a
rural pond receiving animal fecal wastes, Calderon et al. (1991) found swimmer illness to be
correlated with numbers of swimmers and staphylococci densities. They concluded that illnesses
appeared to be caused by swimmer-to-swimmer transmission.
The primary focus of many of the epidemiological studies was gastrointestinal illness
although a few studies examined other health effects (Corbett et al, 1993; Ferley et al., 1989;
Haile etal, 1999; Pike, 1994; Seyfried et al., 1985). Fecal coliforms, Staphylococcus
Aeromonas, Pseudomonas aeruginosa, and viruses were found to be related to a variety of
symptoms as shown in Table 1-11. The studies were conducted in North America, Australia,
and Europe in marine and fresh waters with a variety of discharge types. Fecal coliform was
related to respiratory, ear, eye, and skin illnesses. It was also related to fever and total illness. E.
coli was associated with eye, ear, and skin symptoms. Total Staphylococcus was related to eye,
skin, and total illness. Viruses were related to fever, chills, eye, and respiratory symptoms.
Aeromonas andpseudomonas aeruginosa were related to skin diseases. While these studies
provide general information on relationships between indicators and various ailments, a great
deal of additional research is needed to determine the appropriate microorganisms that should be
measured for assessing risk of nonenteric illness. The organisms are likely to vary for fresh and
marine waters, for different regions of the world, and for different pathogen sources.
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While the use of indicators has these limitations, the use of direct pathogen analysis has
limitations as well as expressed by Pontius and Clancy (2000):
"Current testing methods cannot determine with certainty whether Cryptosporidium
detected in drinking water is alive or whether it can infect humans. In addition, the
current method often requires several days to get results, by which time the tested water
has already been used by the public and is no longer in the community's water pipes.
...Analytical method limitations prevent using Cryptosporidium monitoring data to
accurately assess risk or even to set an acceptable level of risk for Cryptosporidium in
drinking water. Water utilities must be vigilant in applying source water protection and
appropriate treatment to protect customers against this organism."
This scenario can be applied to many pathogens besides Cryptosporidium. Microbial indicators
are not always indicative of pathogens. Since we are ultimately concerned with pathogens, and
not indicators, the relative safe number or infectious dose is often zero. Pathogens are capable of
reproducing to large numbers if only one is originally present. The way to provide a safe
number is to set treatment standards to remove or kill all pathogens that are present
(Warrington, 200la). It therefore is imperative for the development of new water quality
monitoring methods that can identify fecal pathogen contamination, specifically human fecal
contamination (Calderon et a/., 1991). This would provide a greater assessment of the potential
for waterborne disease to occur. Artificial neural networks are modeling tools that have the
potential for predicting peak microbial concentrations and for identifying land use-associated
fecal pollution sources and relative ages of runoff (Brion and Lingireddy, 2003). Using a ratio of
atypical to total coliform colonies (AC/TC) from a membrane filtration analysis of a receiving
water sample, identification of fecal pollution sources has been 90% accurate. Presently, there
are many practices to minimize the exposure to pathogens, such as prudent outfall placement and
maximizing inflow reduction. Chapter 3 discusses management and control of pathogens aimed
at reducing pathogen contribution to receiving waters.
1.5 Conclusions
For water bodies regulated under the CWA and not meeting standards for microbial
contaminants, a TMDL or approach for bringing the water body into compliance must be
developed. All legislation and regulations applicable to the water body need to be considered
throughout the TMDL process. These include the SOW A, CZARA, NPDES permits, and
BEACH program. In most cases, the TMDL will be established for the microbial indicator
involved in the violations of the state ambient water quality standard. However, in some cases, it
is appropriate to develop and meet an alternative narrative standard that addresses the designated
use impairment. These could include situations where a pathogen source is identified as the
cause of waterborne disease outbreaks.
Waterborne disease outbreaks can be caused by pathogenic bacteria, protozoa, viruses,
helminth worms, and fungi. Worms and fungi have not been identified in the waterborne disease
outbreak data compiled by U.S. EPA and the CDC. These reports identify Cryptosporidium
1-37
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parvum, Giardia lamblia, E. coll O157:H7, and Shigella as causative agents for significant
numbers of outbreaks and illnesses from exposure to contaminated recreational waters and
drinking waters from surface sources. In addition, many waterborne disease outbreaks and cases
go unreported, and there are outbreaks where the pathogen responsible is not identified.
Indicator organisms and monitoring programs are limited in their ability to predict
pathogen presence and health risks. However, because of their ability to provide a general
indication of the presence of human fecal material and their low cost and complexity, microbial
indicators have remained the primary means for assessing microbiological contamination in
water. Fecal coliform has historically been the microbiological indicator of choice, but its
presence does not always correlate well with the incidence of disease. Coliforms are being
replaced by the more specific indicators Enterococcus and E. coli for fresh waters, and
Enterococcus for marine waters (U.S. EPA, 2003b). The acceptance of these indicators is
gradually occurring. Once a large database is established for Enterococcus andE. coli, these
organisms will be used by the states with conviction and confidence. Most of the investigations
concerning relationships between total coliform bacteria and pathogenic microorganisms in
environmental waters found poor or no correlations. Of the 13 epidemiological research studies
from around the world reviewed for this publication, Enterococcus density appears to be the
indicator most strongly correlated with gastrointestinal illness among bathers in recreational
waters. E. coli is also related to gastrointestinal illness in a number of studies. Both of these
organisms were found to be related to enteric illness more frequently than fecal coliforms. Five
of the studies reviewed investigated non enteric effects in addition to gastrointestinal. Fecal
coliforms, staphylococcus, Aeromonas, Pseudomonas aeruginosa, and viruses were found to be
related to respiratory, eye, and skin symptoms as well as fevers. Climate, water type, and
pollution sources (i.e., sewage or stormwater runoff) are all factors affecting the ability of an
indicator to be a predictor of pathogenic pollution and illness.
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Chapter Two
Detection Methods and Alternate Indicator Organismsl
2.1 Introduction
Public demand and regulatory requirements compel monitoring for pathogen risks. Such
monitoring requires feasible and accurate detection methods for appropriately selected microbes.
Water quality monitoring in the U.S. is most frequently conducted for bacterial indicators using
the standard membrane filtration or multiple tube fermentation/most probable number methods.
U.S. EPA requires that a Total Maximum Daily Load (TMDL) be developed for water bodies
violating standards, which are determined using the monitoring results. The TMDL is generally
developed for the microorganism responsible for the violation. There are exceptions, however,
such as when there are waterborne disease outbreaks. In these instances, other detection
methods may need to be employed to identify causative agents and determine their presence and
concentrations in a watershed.
Microorganisms responsible for waterborne disease outbreaks are identified through
clinical testing of individuals who seek medical care for their illness. Illnesses are classified as
waterborne disease outbreaks when more than one individual is found to be infected with the
same microbe believed to be from a common source of drinking or recreational water.
Environmental officials assigned to investigate and manage the pollution responsible may use
microbial source tracking and pathogen detection methods to investigate possible sources and
determine the extent of contamination.
This chapter presents information on detection methods for bacteria, viruses, and
protozoa, summarized in Tables 2-1, 2-2, and 2-3, respectively. In the section on bacteria,
detection methods for both indicators and pathogens are discussed, as well as alternatives to the
traditional indicator organisms and an overview of selected methods for microbial source
tracking. Although helminths and fungi are discussed in Chapter 1, their methods were not
reviewed for this chapter due to the high unlikelihood that these organisms will be encountered
in urban watersheds in the U.S. Information about pathogenic fungi is available in Standard
Methods for the Examination of Water and Wastewater (Clesceri et a/., 1998), hereafter referred
to as Standard Methods. A method for helminth ova is presented in the U.S. EPA document
Control of Pathogens and Vector Attraction in Sewage Sludge (U.S. EPA, 1999) available at
http://www.epa.gov/ORD/NRMRL/Pubs/1999/625R92013.pdf.
1 Much of this chapter is excerpted from Monitoring Pathogens in the Watershed: Indicator
Organisms and Detection Methods, submitted for peer-reviewed publication by M.E. Tuccillo and J.M.
Perdek, and Investigating Watershed Microbial Pathogen Contamination to Manage Public Health Risks,
submitted for peer-reviewed publication by J.M. Perdek, M.E. Tuccillo, and S.M. Wankel.
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2.2 Detection Methods
2.2.1 Bacteria
2.2.1.1 Cultural and Enzyme-Based Methods
Cultural methods, or those that grow bacteria in a prepared medium, have been used for
indicator bacteria detection and enumeration for over a century (Pyle et a/., 1995). Membrane
filtration methods are well established and routinely used. The details of these methods are
described in StandardMethods. The water sample is filtered, the filters are incubated on a
growth medium for a specific time and temperature, and the resulting colonies are enumerated.
The membrane filtration incubation period is 24 hours for fecal coliforms and total coliforms, but
for other bacteria can take longer; Staphylococcus aureus and enterococcus cultures must be
incubated for 48 hours, and Pseudomonas cultures should be incubated for 72 hours. An
improved U.S. Environmental Protection Agency (EPA) method for enterococcus using a
modified type of agar (mEI) requires only 24 hours of incubation (U.S. EPA, 1997). The
membrane filtration methods require confirmation tests, which entail further effort and additional
incubation time.
Multiple-tube fermentation/most probable number methods for coliform bacteria are
based on the ability of the organisms to ferment lactose. Tubes with growth medium are
inoculated with a series of undiluted and diluted samples, with several tubes inoculated per
dilution. Following incubation at the specified temperatures, the numbers of tubes
demonstrating a positive response are recorded and a statistical estimate of the bacterial density
is determined. Most probable number methods take 48 hours for coliform incubation, plus an
additional 24-hour confirmation test. Enterococcus and fecal streptococcus are incubated for 24-
48 hours with an additional 24 hours for confirmation (Clesceri etal., 1998). Fecal coliforms,
however, can be analyzed in 24 hours by the A-l broth 1-step method (StandardMethods #
9221E.2).
Methods that rely on counting the colonies that form during incubation, including
membrane filtration, tend to underestimate bacterial numbers (Sartory et a/., 1999). This
phenomenon affects analyses for both indicators and pathogens. This may be due to clumping,
particle association, cell injury, and the viable-but-nonculturable (VBNC) state of the bacteria.
In the VBNC state, cells may maintain viability and metabolic activity, but fail to grow and
multiply on culture plates. Huq and Colwell (1996) reviewed this topic with special attention to
Vibrio cholerae, although this condition applies toAeromonas, Shigella, Staphylococcus, and
Campylobacter, among others. Such underestimation of bacterial counts presents the obvious
danger of giving rise to misleading reports.
Substrate hydrolysis by a specific enzyme with colorimetric endpoints forms the basis of
several detection methods. In substrate hydrolysis, the hydrolysis reaction between an enzyme
in the bacteria and the substrate results in a color change that is used to determine the analytical
results. Cultural methods for E. coli are based on detecting the action of the enzyme |3-
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glucuronidase upon the substrate 4-methylumbelliferyl-|3-D-glucuronide (MUG) (Sartory et al,
1999; Shadix et al., 1991). The product fluoresces blue under long wavelength ultraviolet (UV)
light, indicating the presence of E. coll. The E. coli technique in Standard Methods requires
additional incubation of coliform-positive membrane filtration samples to test for MUG
utilization by P-glucuronidase. The U.S. EPA method using membrane-Thermotolerant E. coli
(mTEC) agar for E. coli analysis (U.S. EPA, 1985) relies upon detection of the enzyme urease; a
modified mTEC method relies upon P-glucuronidase. Substrate hydrolysis by p-galactosidase is
used for detection of thermotolerant coliforms. The rapid method tested by Robertson et al.
(1998) uses only a 6-hour incubation to test for P-glucuronidase for E1. coli and P-galactosidase
for thermotolerant coliforms.
There are rapid alternatives to membrane filtration methods based on enzyme substrate
utilization by coliform bacteria and enterococci. IDEXX Laboratories (Westbrook, ME)
produces a series of widely used EPA-approved products. Their Colilert® Quantitray™ , which
uses their patented Defined Substrate Technology, is an easy-to-use commercial most probable
number method designed for simultaneously determining the presence of total coliforms andE.
coli in 24 hours (Edberg et al., 1989; Townsend et al., 1996). Total coliforms are detected by
the action of P-galactosidase, and E. coli detection is based on the action of P-glucuronidase.
Coliforms produce a yellow product and E. coli produces a product that fluoresces yellow.
Colilert-18® permits detection of these organisms in only 18 hours. Colilert® has been shown
by some researchers to be as sensitive as Multiple Tube Fermentation (MTF) and membrane
filtration (Eckner, 1998; Fricker et al, 1997; Edberg et al, 1990). Francy and Darner (2000)
used recreational water to compare Colilert to the U.S. EPA-recommended mTEC method (U.S.
EPA, 1985), a p-glucuronidase-based membrane filtration technique. The authors found
statistically significant differences between the methods, but note that their test area was small
and further work is needed. The expression of P-glucuronidase can, however, be suppressed by
environmental stress (Sartory et al, 1999; Edberg et al, 1990), raising the possibility of
underestimating bacterial densities. Furthermore, E. coli O157:H7 does not possess this enzyme,
so a separate test for E. coli O157:H7 would be needed if it is suspected.
Similar to Colilert®, IDEXX Laboratories' enzyme-based Enterolert® method is
designed to provide a most probable number method in 24 hours for enterococcus in water. The
hydrolyzation product of the substrate fluoresces blue. Abbott et al (1998) found a positive
correlation between Enterolert® and membrane filtration in marine waters in New Zealand.
Budnick et al (1996) and Eckner (1998) reported equal or better sensitivity and specificity with
Enterolert compared to membrane filtration in recreational waters.
Because indicator bacteria are used as a basis for public health decisions in dynamic
aquatic environments such as beaches, long analysis times are problematic because levels of E.
coli and thermotolerant coliforms fluctuate. Fortunately, there are rapid method alternatives to
the commonly used cultural methods can speed decision-making about protective measures such
as beach closings. The 18-hour incubation time for Colilert and the 6-hour incubation used in
the method of Robertson etal (1998) are two examples of incubation methods that require less
time. In addition to rapid cultural methods, other classes of detection methods, such as
immunological and genetic techniques, offer possibilities for faster analysis times.
2O
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2.2.1.2 Immunological Methods
A group of immunological detection methods for microorganisms is based on the use of
antibodies, which bind with antigens on the organism's surface. A limiting factor with all
immunological techniques is the specificity of the antibody used. Ideally an antibody should
bind only with a single antigen, thereby targeting only the organism of concern. Monoclonal
antibodies (Mabs) are clonally derived from a single antibody-producing cell. This means that
they are exceptionally pure and highly specific in their action.
Some immunological methods are applicable for efficient bacterial detection methods. In
immunofluorescence (IF), the antibodies are tagged with a dye that fluoresces under UV light;
enumeration can be accomplished by epifluorescent microscopy. Sartory and Watkins (1999)
note that there is promise for a limited cultural period (4-6 hours) coupled with detection either
by substrate light emission or immunological techniques for same-day results. In their review of
rapid methods, McFeters etal. (1999) cite examples of the staining of bacteria with fluorescent
antibodies performed directly on membrane filters. This avoids steps such as sample
concentration and fixation on glass slides.
Because pathogenic E. coli O157:H7 does not produce p-glucuronidase, the E. coli
procedures in Standard Methods will not detect it without additional steps. Immunological
techniques may be useful in situations where this pathogen is suspected. The rapid E. coli O157:
H7 methods of Pyle etal. (1995, 1999) involve incubation with a dye that indicates viability,
followed by fluorescent antibody staining and enumeration by epifluorescent microscopy or laser
scanning cytometry (the study and measurement of cells). Kfir etal. (1993), however, caution
against problems of specificity with the use of monoclonal antibodies as a rapid tool for
detecting fecal bacteria in water, and in particular E. coli.
Commercially available instruments such as Chem Scan® can detect and enumerate
fluorescent bacteria (McFeters et al., 1999), further facilitating rapid detection methods.
Commercial sensors continue to be developed and were reviewed by Ivnitski et al. (1999). Some
are immunologically based; others rely on enzyme detection or nucleic acid detection. A rapid
immunological technique for E. coli O157 and Salmonella typhimurium (Yu and Bruno, 1996)
uses a commercial sensor and shows promise as a screening tool, identifying samples that should
be further analyzed. These simplified, commercial screening tools provide additional options for
situations where easy, rapid screening is desired.
A process called enzyme-linked-immunoabsorbent assay (ELISA) tags an antibody with
an enzyme. After incubation, an enzyme substrate is added, and the formation of a pigmented
product is indicative of the amount of enzyme present in the sample and, therefore, the amount
of microorganism in the sample (Bitton, 1980). Advantages of ELISA are that it is robust,
versatile and simple to perform (Kfir et al., 1993). As with any immunoassay, limitations are
related to the specificity of the antibody used. Various easy-to-use commercial ELISA kits are
available, such as the Wellcolex kits (Murex Biotech Dartford, United Kingdom). Developed
mostly for clinical or food applications, these techniques may be useful for water quality testing
when simple techniques are desired. Limited trials with wastewater have, however, raised the
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possibility of cross reactions with competing organisms in the samples (Meckes, 2001). Further
testing of these kits with environmental waters is needed.
2.2.1.3 Genetic Methods (Gene Probes and PCR)
Development of genetic methods has provided new sensitive options for pathogen
detection. Gene probes are nucleotide sequences that pair with corresponding sequences in the
sample through a process called hybridization (Hurst et al, 1989). The probes make good
detection tools and can be labeled with a radioisotope, an enzyme, or a fluorescent chromogene
to permit detection. Although genetic methods require more sophisticated equipment and
techniques than cultural methods, there are commercially available gene probe kits that only
require typical microbiological laboratory equipment and are easy to use. For example, Gene-
Trak (Hopkinton, MA) produces gene probe assays for several organisms including E. coli and
Salmonella. The kits, which are geared primarily toward food or clinical applications, have a
colorimetric endpoint and come with a photometer. Rice et al. (1995) found that the probe
performed well with pure cultures, but failed to detect seven of thirteen positive cultures in creek
and river samples, possibly due to low bacterial densities in the natural waters. The authors note
that further research is needed to improve the performance of the method with environmental
samples, possibly through increased enrichment or larger sample aliquots.
The polymerase chain reaction (PCR) has greatly improved the ability to detect low
densities of pathogens in environmental samples. PCR produces many copies of a target section
of a microorganism's deoxyribonucleic acid (DNA). With the large number of copies produced
by PCR, the target DNA can be detected using gene probes or gel electrophoresis (Toze, 1999).
Gel electrophoresis is a process used to impart an electric current to DNA fragments in a gel of
specific density. Different size fragments move at different rates and can be visualized as a
series of bands in the gel. The use of PCR offers several advantages, including specificity,
sensitivity, rapidity, accuracy, and the capacity to detect small amounts of target nucleic acid in a
sample. PCR-based methods can be used both to rapidly identify bacteria that have been isolated
and for direct pathogen detection in environmental samples (Toze, 1999).
Several researchers have published protocols for PCR-based detection of E. coli in water
(Flicker et al, 1999; Kong et al, 1999b; Tsen et al, 1998). The method of Fricker et al (1999)
is especially quick, identifying E. coli from membrane filters within two hours. Tsen et al
(1998) use an 8-hour pre-culture step, and claim detection of 1 cfu per 100 mL. By combining
PCR and radiolabeled gene probes, Bej et al (1990) developed a sensitive and specific method
for E. coli, Salmonella and Shigella spp. A PCR method for Salmonella spp. published by Way
et al (1993) can also detect other coliform bacteria (e.g., Shigella, E. coli and Citrobacter),
rendering the technique very useful for environmental samples. Palmer et al. (1993) found PCR
to be sensitive and specific for Legionella in sewage treatment plant influent and in ocean
receiving waters. A method for detecting Aeromonas in seawater (Kong et al, 1999a) may be
useful for monitoring because of the prevalence of Aeromonas spp. in the aquatic environment.
There are, nevertheless, several disadvantages to PCR-based methods. They require
specialized equipment and skilled technicians (Toze, 1999). The results of PCR alone do not
2-5
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provide a means for quantification; they indicate presence or absence of the target genetic
material. Furthermore, PCR alone does not directly provide information about the viability or
infectiousness of the organisms because DNA may persist in the environment (Alvarez et al.,
1993; Gantzerera/., 1999; Kopecka etal., 1993; Metcalf etal., 1995; Sobsey e^a/., 1998a).
These techniques are still at the research stage and are beyond the capabilities of most state and
local municipalities for routine analyses. However, they may eventually become a viable option
for routine pathogen analysis and may be especially useful for studies characterizing the
identities and sources of pathogens within a watershed.
2.2.2 Viruses
Current routine monitoring strategies do not test for viruses; they rely on indicator
bacteria. Various viruses (e.g., rotavirus, adenovirus, hepatitis A virus and Norwalk-like
viruses) are important agents of illness in sewage-polluted waters (Metcalf et al, 1995). There
are clearly cases where virus identification is needed, such as in investigations of outbreaks or in
research studies. In cases where direct detection of viruses is needed, a variety of methods exists
and new methods continue to be developed.
2.2.2.1 Sample Concentration
Because of the low concentrations of viruses in environmental samples, methods used to
detect enteric viruses require an initial concentration step to make them detectable. For
environmental waters this is typically accomplished by sorption of viruses onto a filter.
According to Schwab et al. (1993), hundreds to thousands of liters of water may need to be
filtered through a special filter cartridge to achieve sufficient virus concentration for detection.
A yarn fiber filtration cartridge or a cartridge with pleated sheets of filter material are
particularly useful because of field portability. After filtration, the viruses are generally
recovered from the filter into about 1L of eluant. Standard Methods describes techniques for
virus concentration by adsorption to and elution from microporous filters. Beef extract is one of
the most common eluants (DeLeon and Sobsey, 1991; Schwab etal., 1993). A secondary
concentration step may be needed, such as ultrafiltration or flocculation. In their review of
filtration and elution methods, DeLeon and Sobsey (1991) caution that humic and fulvic
substances in water may interfere with virus sorption onto filters. They also point out that
adsorption/elution efficiencies vary for different viruses; for some the recoveries are low.
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Table 2-1. Summary of Detection Methods for Bacteria
Cultural and Enzyme-Based
Method
Membrane Filtration
Multiple Tube
Fermentation/Most
Probable Number
(MTF/MPN)
Substrate Hydrolysis -
Colorimetric
Defined Substrate
Technology
Duration
24 hours or longer
depending on bacteria +
24-hour confirmation
24 hours or longer
depending on bacteria +
24-hour confirmation
6 to > 24 hours
depending on method
and organism
£. co/; and Total Coliform
- 24 hours; Enterococcus
- 24 hours
Results Provided
Enumeration,
Presence-Absence
Enumeration,
Presence-Absence
Presence-Absence
Enumeration
Capabilities Needed
General Microbiology
Laboratory
General Microbiology
Laboratory
General Microbiology
Laboratory
General Microbiology
Laboratory
Immunological
Immunofluorescence (IF)
Commercially Available
Instruments
Enzyme-Linked-
Immunoabsorbent Assay
(ELISA)
< 24 hours
< 24 hours
Varies
Enumeration by
epifluorescent
microscopy
Enumeration
Enumeration
Specialized
Microbiology Lab.
General Microbiology
Laboratory
Kits available for
clinical and food
applications; more
research for
environmental app.
needed
Genetic
Gene Probes
PCR
Time varies
< 24 hours
Presence-Absence,
Enumeration in
research stage
Presence-Absence,
Enumeration in
research stage
Kits available for
clinical and food
applications; more
research for
environmental app.
needed
Specialized
Microbiology Lab.;
techniques still in
research stage
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Table 2-2. Summary of Detection Methods for Viruses
Method
Duration
Results Provided
Capabilities Needed
Cultural
Cultural Assay
Varies, on the order
of days
Presence-Absence,
enumeration; indicates
viability
General Microbiology
Laboratory
Immunological
Immunological
Immunological:
ELISA
Varies
Varies
Enumeration by
epifluorescent microscopy;
Does not indicate viability
Presence-Absence;
Enumeration
Specialized
Microbiology Lab.
More research
needed for
environmental app.
Genetic
Gene Probes
PCR
Varies
< 24 hours
Presence-Absence by
radioisotope or enzyme
Presence-Absence
Specialized
Microbiology Lab.;
More research for
environmental app.
needed
Specialized
Microbiology Lab.;
techniques still in
research stage
2.2.2.2 Cultural Assay
Several assay techniques are available for virus detection in the concentrated sample.
Detection methods given in Standard Methods rely on the infection and destruction of host cells
by the virus (cytopathic effects). In the plaque assay method, for example, a viral suspension is
placed on a monolayer of cells, and areas of cell destruction due to infection (plaques) are
enumerated and expressed as plaque-forming units (PFU). An advantage of cell culture is that it
indicates viability. There are, however, disadvantages. Cell culture assays such as the plaque
assay method require different cell lines for detection of different viruses. Although most enteric
viruses can be cultured, some viruses, such as Norwalk virus, hepatitis A and E, calciviruses,
rotaviruses and astroviruses either do not grow or grow slowly in cell culture assays (DeLeon
and Sobsey, 1991; Metcalf et al., 1995). Thus, cell cultures cannot be used to detect several
important pathogenic viruses.
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2.2.2.3 Immunological Techniques
Immunological techniques are useful in virus detection. The viruses may be in
suspensions, trapped on filters, or in cell cultures (Hurst et al., 1989). When they are trapped on
filters, without some form of cell culture, the assay cannot indicate infectivity. As with the
bacterial techniques mentioned earlier, use of a fluorescently-tagged antibody permits
enumeration by epifluorescent microscopy. Oragui et al. (1989) have used immunofluorescence
for detection of rotaviruses in wastewaters. Similarly, radioimmunoassay uses an antibody
tagged with a radioactive isotope to bind to the viral antigen, and detection is accomplished by
measuring the radioactivity of the antibody-antigen complex.
Variants of the enzyme-linked assay can detect viral antigens trapped on a filter or
associated with infected cells (for viruses that can be cultured). ELISA has been used to detect
Hepatitis A virus in tap water (Schnattinger, 1985). Nasser and Metcalf (1987) and Nasser et al.
(1993) developed an amplified ELISA (A-ELISA) method for virus detection that has greater
sensitivity than ordinary ELISA, as well as good specificity, speed, and low cost. Nasser et al.
(1994) used A-ELISA to indicate the presence of viable poliovirus in water. According to Kfir
and Genthe (1995), commercial clinical ELISA kits have been used for environmental waters
and are available for some viruses, including rotaviruses and adenoviruses.
2.2.2.4 Gene Probes
Viruses may be detected by the use of gene probes. As with the immunological methods,
the target material may be present in a solution, trapped on a filter, or present in infected cells.
Detection may be accomplished via a radioisotope or enzyme attached to the gene probe. An
effective method must specify a target nucleic acid sequence that is specific to the organism of
concern. As with other assays, prior amplification by cell culture indicates that the viruses are
infective. Hurst et al. (1989) note that hybridization is more sensitive and faster than plaque
assays or immunofluorescence. According to Gerba et al. (1989), hybridization is much more
sensitive than ELISA methods, and gene probes have been developed for the major groups of
enteric viruses. Gene probes have been used for the detection of hepatitis A virus and other
enteroviruses in drinking water samples that were negative by radioimmunoassay and that
required weeks of propagation in cell cultures to be detectable by immunoassays (Shieh etal.,
1991). Other examples of studies using gene probes include the detection of rotavirus in fresh
and estuarine waters (Nasser et al., 1991), enteric viruses in raw and treated waters (Genthe et
al., 1995), and poliovirus in sewage-contaminated groundwater (Margolin et al., 1990).
Margolin et al. (1993) found excellent agreement between cell culture and gene probe methods
for a variety of environmental water samples. As noted earlier, however, genetic techniques
require sophisticated equipment and techniques. The research studies show promise for efficient
viral detection, but easy-to-use kits are not readily available.
2.2.2.5 PCR-based Methods
The polymerase chain reaction is particularly useful for virus detection because it
amplifies the low quantities of viral genetic material present in environmental samples. The use
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of PCR for detecting viruses offers many advantages over the traditional methods, including
lower detection limits, increased range of viruses detectable, specificity, and shorter processing
time (Toze, 1999). As with other methods, water samples may need to be filtered or otherwise
concentrated first. Reverse transcriptase, a compound that catalyzes the formulation of DNA
using RNA as a template (RT-PCR), is used when a virus' genetic material is RNA. The RT-
PCR methods can detect less than 10 PFU of a virus in a filter eluate sample in less than two
days.
Standard sample concentration procedures can pose problems for PCR. Humic acids,
which cause interference, can be concentrated along with the viruses. Proteins and salts in beef
extract eluant can also interfere with molecular methods (Schwab et a/., 1993). It is, therefore,
necessary to separate the viruses and their DNA from such impurities (Kopecka et a/., 1993).
The inhibitory problems in some samples have been avoided by using immunologic-based
methods to capture viruses for subsequent PCR amplification (Metcalf etal., 1995; Schwab et
a/., 1996; Toze, 1999).
Polymerase chain reaction-based techniques have been used successfully for detection of
viruses in various types of environmental samples, often with relatively short analysis times.
Methods have been developed for astroviruses (Marx etal., 1998), enteroviruses (Gilgen etal.,
1995; Griffin etal., 1999; Vantarakis and Papapetropoulou; 1998, 1999), rotaviruses (Soule et
a/., 2000), and adenoviruses (Vantarakis and Papapetropoulou, 1998, 1999) in a variety of
environmental waters. In a comparison of three detection methods for enteroviruses in activated
sludge and sewage waters, Kopecka et al. (1993) found PCR to be vastly more sensitive than cell
culture methods and direct hybridization. A number of RT-PCR methods offering various
advantages have been devised. These include a triple RT-PCR method for the simultaneous
detection of hepatitis A virus, poliovirus, and rotavirus (Tsai etal., 1994), an assay for
enteroviruses with a tissue culture state to indicate infectivity (Fricker et a/., 1999), and a
relatively rapid method using RT-PCR, followed by hybridization and a form of ELISA
(Greening etal, 1999).
2.2.3 Cryptosporidium and Giardia
2.2.3.1 Immunofluorescence
As with viruses, identification of Cryptosporidium parvum oocysts in water is not
routine, limiting our ability to assess the public health threat from Cryptosporidium (Rose,
1997). The public health impacts of this organism are discussed in detail in Chapter 1. The
detection procedure for Cryptosporidium parvum oocysts and Giardia lamblia cysts described in
Standard Methods is an immunofluorescence (IF) procedure. To prepare the sample, hundreds
of liters of water are passed through a filter cartridge. Cysts and oocysts are recovered from the
cartridge, concentrated, and filtered onto a membrane. In addition to the epifluorescent
microscopy phase, contrast microscopy is used for confirmation of the internal structures of the
organisms. The newest U.S. EPA-recognized IF method for Cryptosporidium and Giardia (U.S.
EPA, 2001) is a more streamlined method that entails filtration of only 10 L of water, uses well
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slides instead of membrane filters, and uses differential interference contrast (DIG) microscopy
for confirmation.
The IF procedures have low recoveries, are costly and time-consuming, and cannot
indicate viability (Slifko et al., 1997). The most recent edition of Standard Methods
acknowledges these limitations, but does not provide an updated method, noting that methods
research is evolving rapidly. Allen et al. (2000) note that IF techniques have a high rate of both
false positives and false negatives, rendering monitoring results highly suspect.
Two methodologies address the problem of viability. Jarmey-Swan et al. (2000)
improved upon IF for Giardia cysts by staining with fluorescein diacetate prior to antibody
staining. The combination of the two stains allows identification of viable cysts via microscope.
Slifko etal. (1997, 1999) have developed and statistically standardized a detection method based
on cell culture technology combined with an IF assay. The technique, called the Foci Detection
Method (FDM), can be used to detect concentrations as low as 10 oocysts per sample. This
method has good promise of being a specific test for Cryptosporidiumparvum, but it has not yet
been tested with all Cryptosporidium species.
Table 2-3. Summary of Detection Methods for Cryptosporidium and Giardia
Method
Duration
Results Provided
Capabilities Needed
Immunological
Immunofluorescence
72-96 hours
Enumeration by
epifluorescent and contrast
microscopy; Does not
indicate viability
Specialized
Microbiology Lab.
Genetic
Gene Probes
PCR
Time varies
< 24 hours
Presence-Absence
Presence-Absence; does not
indicate viability
Specialized
Microbiology Lab.;
more research for
environmental app.
needed
Specialized
Microbiology Lab.;
techniques still in
research stage
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2.2.3.2 Gene Probes and PCR-Based Methods
While immunofluorescence remains the primary approach for Giardia and
Cryptosporidium analyses, work is continually underway to devise improved techniques that
may replace the current methods. Rose (1997) notes that PCR, ELISA, cultural,
immunomagnetic separation (IMS), and colorimetric methods are not yet sufficiently developed
for routine use. Below is an overview of methods employed in research studies; these may point
the way for future routine detection options.
As an alternative to the antibody approaches, gene probes have been used with
fluorescent staining of Cryptosporidium parvum oocysts in water (Vesey et al., 1998). Prescott
et al. (1999) describe the use of gene probes for the detection of Cryptosporidium parvum. The
method has good specificity and determines viability.
Studies using PCR for detection of Cryptosporidium and Giardia (Rochelle et al., 1997;
Stinear et al., 1996; Ware et al., 1995) have shown that PCR has excellent sensitivity.
Furthermore, simultaneous detection of Cryptosporidium and Giardia is possible. Wiedenmann
et al. (1998) provide a thorough review of PCR for the detection of Cryptosporidium parvum.
As with viruses, methods are available for separation of cysts and oocysts from substances that
can inhibit PCR. For example, a technique called the Xtra Bind Capture System has been used
to facilitate the concentration of Cryptosporidium from water prior to RT-PCR (Kozwich et al.,
2000). In this method, potential inhibiting contaminants were removed and PCR amplification
was performed without needing to elute the oocysts from the capture material. The authors
completed the analysis within only three hours. Other rapid and sensitive PCR methods combine
immunomagnetic (magnetic beads with antibodies) separation of Cryptosporidium oocysts,
followed by PCR for amplification and hybridization for detection (Hallier-Soulier and Guillot,
1999; U.S. EPA 2001). Champliaud et al. (1998), however, note difficulties differentiating
between Cryptosporidium parvum and other nonpathogenic Cryptosporidium species using PCR.
Furthermore, as with viruses, PCR alone cannot indicate protozoan viability. An alternative is to
use messenger RNA (mRNA) for the PCR. The mRNA tends to have a short half life and
therefore should not be present to be recovered from dead organisms (Wiedenmann et al., 1998).
2.3 Alternative Indicator Organisms
2.3.1 Clostridium perfringem
Clostridium perfringens is a hardy, spore-forming bacterium that has potential use as an
indicator of pathogenic bacteria, viruses, and protozoa. In wastewater treatment and disinfection
evaluations, C. perfringens was found to be more disinfection-resistant than fecal coliform and
enterococcus, and was a good indicator of the inactivation of Cryptosporidium parvum oocysts
(Sobsey et al., 1998b). It was also found to be a good indicator for human enteric viruses,
Cryptosporidium, and Giardia in treated drinking water and river water (Payment and Franco,
1993). Research by Kueh et al. (1995) demonstrated correlations between gastrointestinal
symptoms and concentrations of Clostridium perfringens. In marine waters it has been found to
correlate with Salmonella spp. (Morinigo et al., 1992) and Giardia andAeromonas densities
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(Ferguson etal, 1996). C. perfringens has several desirable characteristics, including its
presence in human feces but not bird droppings, and the superiority of spore survival to human
pathogen survival. Furthermore, it can be easily and reliably enumerated using a membrane
filter method.
2.3.2 Bacteriophages
Bacteriophages, viruses that infect bacteria, show promise as water quality indicators.
Almost all bacteria known today have one or a group of specific bacteriophages that infect them.
Coliphages are bacteriophages specific to coliform bacteria. As with C. perfringens, coliphages
were found to be more resistant to disinfection than E. coli, fecal coliform and enterococcus in
evaluations of wastewater treatment and chlorine disinfection (Farrah et al., 1993; Sobsey et al,
1998b).
Bacteriophages that infect through the bacterium's pili are called F+ (male-specific)
phages, and bacteriophages that infect through the bacterium's membrane are called somatic
phages. Studies have found F+ bacteriophages to be effective indicators of enteric virus
concentrations in fresh waters (Havelaar et al., 1993; Nasser and Oman, 1999). Lucena et al.
(1996) suggested using phages of Bacteriodes fragilis, C. perfringens, and sometimes
enteroviruses as indicators of persistent fecal pollution in marine sediments. In an urban
estuarine study, however, F+ RNA bacteriophages did not correlate well with the pathogens
measured (Ferguson et al., 1996). Serrano et al. (1998) found that F+ RNA phages had low
correlations with microbiological parameters in coastal waters, but that coliphages had
statistically significant correlations with microbiological parameters. More evaluations are
needed before a consensus will be reached regarding the selection and use of bacteriophages as
indicators in various types of receiving waters.
2.4 Microbial Source Tracking
Attempts to reduce loads and prevent outbreaks via watershed management can be aided
by accurate determination of the sources of microbial contamination. Microbial source tracking
(MST) techniques can help give an indication of whether the sources of indicators or pathogens
are human, wildlife, or agricultural. Categories of MST techniques include, among others,
phenotypic and genetic methods, and may or may not require the development of a library of
known samples for comparison with unknown samples. Drawbacks for MST methods include
uncertainty in the spatial and temporal stabilities and variabilities of target characteristics. Ease
of use and costs are also important in determining whether a method can be widely applied.
While a summary is provided here, a critical review conducted by fellow EPA researchers
(Simpson et al., 2002) can be reviewed for more detailed information.
2.4.1 A ntibiotic Resistance A nalysis
Antibiotic resistance analysis (ARA) is a phenotypic method that takes advantage of the
exposure of bacterial sources to different antibiotics and the resulting patterns of resistance that
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develop. To determine a multiple antibiotic resistance (MAR) profile, a bacterial isolate is
exposed to a suite of antibiotics. The antibiotics to which the isolate is resistant define the MAR
profile, which acts as a fingerprint. First, a database of MAR profiles is acquired for samples of
known sources in a given region. MAR profiles of unknown samples can then be compared to
the database to determine their probable sources.
Wiggins (1996) analyzed 1,435 fecal streptococci isolates from animal and human
sources for their resistance to five antibiotics. He then used discriminant analysis of the
resulting patterns to classify the known isolates with a high rate of correct classification (92% of
human isolates). Parveen et al. (1997) used MAR profiles to investigate E. coli sources within
Apalachicola Bay and were able to identify MAR profile differences between point and nonpoint
sources. Hagedorn et al. (1999) used antibiotic resistance in fecal streptococci to identify
sources of nonpoint fecal pollution. Antibiotic resistance patterns have also been used in
subtropical surface waters (Harwood et al., 2000) and industrially perturbed stream waters
(McArthur and Tuckfield, 2000). The analytical techniques for obtaining an antibiotic resistance
profile are easy to perform. Antibiotic resistance patterns are, however, region-specific and
compiling a MAR database of known sources is labor intensive. Furthermore, the MAR profiles
of bacterial populations may shift with time. This approach may be best used in small
watersheds with demonstrated nonpoint source problems and a limited number of potential
sources (Simpson et al., 2002).
2.4.2 Molecular Methods
The advance of molecular-based methods in recent years has aided source identification
through the use of genetic markers. More commonly applied to microbial indicators because of
their prevalence in the environment, these molecular-based MST methods are an active area of
research and development. The review prepared by Simpson et al. (2002) describes the state of
development of a number of techniques as well as their advantages and drawbacks. The genetic
methods described in the review include ribotyping, length heterogeneity-PCR (LH-PCR),
repetitive PCR (REP-PCR), denaturing gradient gel electrophoresis (DGGE), pulsed-field gel
electrophoresis (PFGE), and amplified fragment length polymorphism (FLP). Although not yet
ready for routine use, genetic methods are being tested in research studies. For example, a
library-dependent PFGE was used to identify coliform sources in Northern Virginia's Four Mile
Run Watershed (Simmons et al., 2000). The study concluded that nonhuman species (waterfowl,
raccoon, dog, deer, and Norway rat) were the primary E. coli sources in the urban stream.
Human sources contributed only 18% of the E. coli (NVRC, 2002).
Because of the lack of a therapeutic cure or drug therapy for cryptosporidiosis, MST
techniques for Crytosporidium parvum oocysts are particularly appealing. The Centers for
Disease Control (CDC) has evaluated a molecular species- and strain-specific method for
analyzing Cryptosporidium parasites in environmental samples (Royer etal., 2002; Xiao etal.,
2000; Xiao et al., 2001). The method is a nested PCR-restriction fragment length polymorphism
technique. It produces numerous copies of a targeted DNA sequence, uses an enzyme to break it
into fragments and uses gel electrophoresis and staining to separate and visualize the fragments.
Numerous Cryptosporidium species have been examined using this method. It has been tested
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on stream water, surface water, and wastewater, and is claimed to be able to differentiate
between potential sources such as humans, cattle, pets, and wildlife.
In storm stream flow in a mostly undeveloped and forested portion of the New York City
watershed, the procedure identified no genotypes from humans or farm animals, indicating the
genotypes were likely from wildlife. In raw surface water collected less than a mile downstream
of a large commercial cattle operation and a wastewater treatment plant, the method confirmed
the presence of C. parvum human and bovine genotypes. In Milwaukee, wastewater containing
pretreated effluent from a large cattle slaughterhouse was found to contain several genotypes that
were known to be associated with humans, bovines, dogs, cattle, and rodents. The method used
by CDC to identify Cryptosporidium sources shows promise, but needs further development
technologically and is as yet too expensive for routine monitoring (Xiao et a/., 2002; Royer et
a/., 2002).
2.5 Conclusions
Speed, reasonable cost, accuracy, and the level of difficulty in performing the techniques
remain considerations in the selection and execution of microbiological analyses for water
quality. For analysis of total coliform, fecal coliform, enterococcus, andE1. coli, membrane
filtration methods are well established and straightforward to perform without specialized
equipment. Disadvantages include length of analysis times and potential underestimation.
Rapid commercial enzyme-based methods such as Colilert® and Enterolert® show promise for
easy screening. This is especially useful in situations where water quality can change rapidly,
requiring frequent testing. Users should initially test rapid methods against the traditional
membrane filtration or most probable number techniques in order to check their technique and
understand any limitations of the methods. Because E. coli O157:H7 lacks the enzyme |3-
glucuronidase, a separate test, such as an immunological method, is needed if its presence is
suspected. Commercial gene probe kits are available for some bacteria such as E. coli and
Salmonella. Commercial ELISA kits can also be purchased. These have been developed for
food and clinical applications; their use for environmental samples can be explored.
Immunofluorescence and ELISA methods are currently available options for detection of
nonculturable viruses and bacteria as well as Cryptosporidium, Giardia, and E. coli O157:H7.
Commercially prepared ELISA kits are available for some viruses. Although not as sensitive as
PCR-based techniques, immunological methods permit quantification. Allen et al. (2000) have
warned, however, of limitations of the IF methods for Cryptosporidium and Giardia, including
poor recoveries and inability to determine viability. Poor recoveries are an issue for viruses as
well because elution efficiencies from filters can be low. Recovery may be less of an issue in the
detection of bacteria, especially indicator bacteria, because they do not need to be retained and
eluted from a filter for concentration. However, recovery and enumeration of pathogenic
bacteria remains an issue when concentrations are low and exposure is high.
Problems with low viral and protozoan concentrations are being overcome by the high
sensitivities of nucleic acid techniques, which include gene probes for detection and PCR for
amplification of small amounts of a pathogen's DNA or RNA. The large number of research
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studies using PCR in the detection of pathogens illustrates the versatility and promise of these
methods. In particular, the ability to detect low concentrations is beneficial because of the low
infectious doses of protozoa and viruses. PCR also permits detection of nonculturable viruses
and viable but nonculturable bacteria. These methods are still at the research stage and they are
not widely available, although they may be in the future. A major drawback to PCR-based
methods is the inability to indicate viability; results should be considered evidence of recent
contamination and should not necessarily imply risk. Expensive and specialized analytical needs
are another drawback.
Although the ability to detect low concentrations of pathogens offers advantages in
pathogen monitoring, results must be interpreted with care. The calculation of pathogen density
from the analysis of a water sample is based on the assumption that the pathogens are distributed
evenly in the water body being sampled. If this assumption is not true, then the absence of
microorganisms in a sample may not mean that the organism is absent in the water. On the other
hand, detection of a pathogen may give rise to an erroneously high estimate of pathogen density
(Allen et al, 2000). Furthermore, pathogen contamination may be transient and easily missed.
Ongoing background sampling is important for establishing the normal microbiological
conditions of a watershed; sampling should also be conducted when a disturbance such as a
storm increases the likelihood of pathogen presence.
Detection methods are continually evolving, but direct routine monitoring for pathogens
is not feasible at this time. Indicator use is far from ideal, but it still represents the most viable
option for a basic level of water quality monitoring. Unfortunately, indicator bacteria make poor
proxies for viruses and protozoa because their survival characteristics are different from those of
viruses and protozoa. Potential incorporation of C. perfringens and bacteriophages into
monitoring strategies may improve the representativeness of the indicator organisms. Because
organisms such as Aeromonas, an opportunistic pathogen, and some fecal coliform have
nonhuman sources, looking only for human-based fecal contamination does not cover all risk
factors. MST techniques can allow watershed managers to determine whether the sources of
indicator or pathogens are human, wildlife, or from domesticated animals. ARA is currently the
easiest to execute, but in time genetic methods may play an increasing role in tracking down the
microbiological sources of water quality impairments.
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Chapter Three
Management and Control of Pathogens
3.1 Introduction
This chapter presents technical information supporting the Total Maximum Daily Load
(TMDL) process for pathogens, specifically Step 5 - Allocations (U.S. EPA, 2001). The
allocations step's objective (U.S. EPA, 2001) is to:
Using total assimilative capacity developed in the linkage
component, develop recommendations for the allocation of loads
among the various point and nonpoint sources, while accounting for
uncertainties in the analyses (i.e., margin of safety) and, in some
cases, a reserve for future loadings.
The information provided will assist watershed managers in determining the capabilities of
control technologies, i.e., disinfection, and best management practices (BMPs) for reducing
microbial concentrations in point and nonpoint sources.
Following is the definition of point sources as presented in the Clean Water Act (CWA),
Section 502 (14):
The term "point source" means any discernible, confined and discrete conveyance,
including but not limited to any pipe, ditch, channel, tunnel, conduit, well, discrete
fissure, container, rolling stock, concentrated animal feeding operation, or vessel or other
floating craft, from which pollutants are or may be discharged. This term does not
include agricultural stormwater discharges and return flows from irrigated agriculture.
Wet weather flows (WWFs) regulated by the National Permit Discharge Elimination System
(NPDES) program are considered point sources. These include combined sewer overflows
(CSOs), stormwater associated with industrial activity, construction-related runoff, and
discharges from municipal separate storm sewer systems (MS4s). MS4 stormwater types
regulated through NPDES permits are described in Section 1.3.1.2. The CWA does not provide
a detailed definition of nonpoint sources; these are defined by exclusion, i.e., anything not
considered a point source in the CWA or EPA regulations. All nonpoint sources are caused by
runoff of precipitation over or through the ground. Therefore, WWFs not covered through
NPDES permits are nonpoint sources (U.S. EPA, 2003a).
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As discussed in Chapter 1, there are detailed procedures and different approaches for
determining recommended loads among the various point and nonpoint sources, while reserving
a margin of safety and room for future loading increases. The TMDL definition is provided in
Section 1.3.1.1. Development of a single waste load allocation (WLA) for all point sources of
pathogens - publicly owned treatment works (POTW) or wastewater treatment plant (WWTP)
effluents, and CSO, sanitary sewer overflow (SSO), and stormwater discharges - within one or
more municipalities in a given urban watershed requires knowledge of treatment system
capabilities and effective control strategies. Different approaches can be used to develop WLAs.
One control strategy, a direct approach, is to calculate respective WLAs because treatment
system capabilities and effective control strategies can be fully quantified. Another control
strategy can be to sum up all the major sources of pathogen discharges. This approach provides
the flexibility of adjusting the proportion of flow and loadings among any of the sources present,
such as stormwater, CSO, SSO, and POTW or WWTP discharge locations, to maximize the
treatment of sewage and load reductions. Point sources are generally discharged from a discrete
point and are treated by control technologies and structural BMPs. However, there are discrete
end-of-pipe or drainage channel conduit discharges that do not fit within the legal definition of a
point source.1
Load allocations (LAs) consist of nonpoint sources and a natural background level of a
given water body. WLAs and LAs pertaining to stormwater, CSO, and SSO occur
intermittently as their origins are WWF events. Therefore, in establishing TMDLs, there needs
to be a conversion of these intermittent loads into daily loads. Also, if there are known
occurrences of untreated CSO and SSO discharges, these should be dealt with and accounted for
independently.
LAs are established for nonpoint sources and, where necessary, may include
implementation of BMPs and source reduction strategies. Discrete discharges and diffuse
sources considered legally to be nonpoint sources can be managed using either control
technologies or BMPs. Diffuse sources are generally managed through nonstructural BMPs.
BMPs will be described in the latter part of this chapter (Section 3.3). In some cases, states have
certain mandatory BMP requirements for specific land activities associated with large fecal
indicator loads, such as confined animal operations or with flood control. Often, implementation
of BMPs occurs through voluntary or incentive programs. Therefore, when establishing nonpoint
source allocations within a TMDL, the documentation should include a reasonable assurance that
the BMP(s) will be implemented and maintained and that the effectiveness of the BMP will be
demonstrated. If pathogen loadings are to be reduced by a BMP, the TMDL strategy may
require a long-term water quality monitoring program to demonstrate effectiveness of the BMP
used.
1 The reader should be aware that "point source" is a legal term, as defined on page 3-1.
It is also commonly used to describe all discrete discharges.
3-2
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The effectiveness of BMPs for controlling stressors in general, and pathogens, in
particular, has not been fully established. There are few references with quantified pathogen
removals. There is a difference between a treatment technology and a BMP (see Table 3-1).
Table 3-1. Distinction between a Treatment Technology
and a BMP for Pathogen Control
Source treated
Effectiveness
Prediction of results
Design
Improvement
Cost
Treatment Technology
(Disinfection for
Pathogens)
discrete end-of-pipe or
drainage channel conduit
discharges
known
reasonably accurate
specific
to the level required
known
BMP
discrete end-of-pipe or
drainage channel conduit
discharges; diffuse sources
uncertain; little data
uncertain
specific
uncertain
known
While the effectiveness and pollutant load reduction by a given BMP may be just an estimate,
the effectiveness of a given technology is usually known and treatment results can be predicted
with reasonable certainty. Although some structural BMPs can perform like treatment
technologies, any misjudgement of treatment effectiveness will either reduce its usefulness
and/or increase costs (Field, 1996).
The common practice for managing stormwater has been the use of structural and
nonstructural BMPs. BMPs can achieve significant environmental improvements, such as
reduction of flow volume and removal of suspended solids by sedimentation and filtration.
BMPs achieve different degrees of removal of toxic substances and nutrients associated with the
removed flow and solids. Removal of pathogens through the use of BMPs can also be associated
with reduced flow and removed solids. Disinfection using treatment technologies is feasible for
stormwater that can be collected and confined, but it is seldom implemented.
The following are three examples of collecting and treating stormwater or dry weather
urban runoff:
1. The city of New Orleans, LA evaluated a prototype disinfection facility for stormwater
using sodium hypochlorite in the late 1960s and early 1970s (Pavia and Powell, 1968);
but they did not implement the practice permanently.
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2. Santa Monica's urban runoff recycling facility (SMURRF, 2000) is treating dry weather
runoff and some wet weather runoff since December 2000.
(http://Epwm.Santa-Monica.Org/Epwm/Smurrf.html).
3. Moonlight Beach urban runoff treatment facility in the City of Encinitas, CA has been
treating dry season urban runoff since September 2002 (Rasmus and Weldon, 2003).
3.2 Disinfection Technologies for Control of Pathogens
3.2.1 Introduction
As long as satisfactory levels of suspended solids concentration and particle size are
achieved upstream, disinfection technologies can achieve effective reduction of pathogen-
contaminated concentrated sources such as:
1. POTW or WWTP effluent; sometimes referred to as secondary effluent
2. CSO, SSO, and stormwater discharges, all referred to as WWF because these discharges
occur during wet weather events
3. Industrial wastewater discharges
4. Confined animal feeding operations (CAFOs)
While disinfection of WWTP effluent (or secondary effluent) and of industrial
wastewater discharges is an established practice (U.S. EPA, 1986a), achieving disinfection of
WWF is difficult. Because WWF is a significant contributor of microbial contamination to
receiving waters, disinfection of WWF released as point sources is warranted.
As stated above, WWF point sources consist of CSO, SSO, and stormwater. Stormwater
draining directly into a receiving water body, rather than through a sewerage system, also falls
under the definition of WWF and may be considered to be either a point source or a nonpoint
source. Human fecal contamination is the main concern for sanitary sewer systems. For
stormwater systems, nonhuman-origin (other warm-blooded organisms) and human-origin fecal
coliform microbial contamination from unauthorized sanitary sewage cross-connections are the
concerns. In combined sewer and storm drainage systems, fecal contamination of both human-
and non-human origin are of concern.
Since issuance of the National CSO Control Policy (U.S. EPA, 1994), which requires
disinfection of CSO after primary clarification, the CSO became the most frequently disinfected
component of WWF. Most WWF disinfection studies, with conventional and alternative
technologies, have been conducted on CSO (U.S. EPA, 2002a). However, all components of
WWF, such as SSO and stormwater, carry significant loads of fecal and pathogen contamination
that would be reduced by disinfection.
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Numerous factors need to be considered in discussing WWF disinfection:
1. Disinfection effectiveness as demonstrated by the pathogen reduction levels
2. The need for a high-rate disinfection process
3. The need for suspended solids removal prior to disinfection
4. A description of individual disinfection technologies in the diminishing order of their
commercial availability for WWF treatment and their relative costs
5. A description of disinfection studies and implementation examples
3.2.2 WWF Disinfection Effectiveness
Disinfection effectiveness is conventionally judged by the reduction of microorganism
densities, generally a bacteriological indicator. Disinfection technologies have been tested using
a variety of bacterial and viral indicators and selected individual pathogenic organisms as well.
Where available, this information is presented in the subsequent sections on individual
disinfection technologies. Different indicators may respond very differently to the disinfection
process. A study by the Massachusetts Water Resources Authority, Boston, MA compared
Enterococcus to fecal coliform data in secondary treated effluent and in effluent from CSO
facilities. The investigators found significant differences between how the indicators respond to
treatment. Satisfactory reduction of fecal coliform was achieved with chlorination, but the
reduction of Enterococcus was unsatisfactory (Rex, 2000).
Development of bacteriological indicators was necessitated by the fact that it is both
impractical and expensive to isolate and measure specific pathogenic organisms in water. Use of
the various indicators is discussed in Chapter 1 and summarized here. A group of enteric
bacteria known as coliform are plentiful in human wastes and easy to measure. Therefore,
bacteria of the total coliform group became the generally accepted indicator for fecal pollution,
even though this group includes different genera that do not all originate from fecal wastes (e.g.,
Citrobacter, Klebsiella, and Enterobacter). An improvement over the total coliform indicator is
the more selective fecal coliform indicator, since fecal contamination of human origin is known
to cause diseases in humans. Fecal coliform selects primarily for Klebsiella and Escherichia coli
(E. coli) bacteria. E. coli is the bacterium of interest because it is a consistent inhabitant of the
intestinal tract of humans and other warm-blooded animals. However, the fecal coliform test is
still not fully specific to enteric bacteria and human-enteric bacteria in particular (O'Shea and
Field, 1992).
As discussed in Chapter 1, stormwater runoff can contain high densities of the non-
human indicator bacteria, and epidemiological studies of recreational waters receiving
stormwater runoff have found little correlation between fecal coliform indicator densities and
swimming-related illnesses (U.S. EPA, 1984; Calderon etal, 1991). In 1986, U.S. EPA
recommended that states begin the transition process to the E. coli and enterococci indicators
(U.S. EPA, 1986b). However, many states still retain the total and fecal coliform criteria. The
most widely used bacteriological criterion in the U.S. is a maximum of 200 fecal coliform/100
mL in waters designated for swimming (Field, 1990). Because the fecal coliform indicator is the
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most widely used, disinfection effectiveness is often reported as reduction of this indicator.
Untreated WWF may contain densities of 1><105 to IxlO7fecal coliform/100 mL. Achieving
hundreds (102) of fecal coliform/100 mL in treated WWF with the use of a given disinfection
technology would indicate a very successful treatment. Achieving thousands (103) of fecal
coliform/100 mL in treated WWF with the use of a given disinfection technology may still
indicate an adequate treatment if there will be a significant dilution upon discharge of the treated
effluent.
3.2.3 Requirement for a High-Rate Disinfection Process
Experience has shown that the long contact time required for conventional wastewater
treatment is extremely costly for the treatment of WWFs due to their relatively high flow rates
and intermittently occuring volumes. However, WWF disinfection can be achieved at shorter
contact times. (U.S. EPA, 1979a; U.S. EPA, 1979b; Stinson etal, 1998). This approach has
been termed "high-rate disinfection." There is currently no clear definition as to what constitutes
high-rate disinfection other than achieving the required bacterial reductions at detention times
significantly less than 30 minutes, the standard contact time (U.S. EPA, 1993).
High-rate disinfection is accomplished by: (1) increased mixing intensity, (2) use of
higher concentrations of disinfectant, (3) use of chemicals or irradiation with higher oxidizing
rates or microorganism-kill potential, or (4) combinations of these (Field, 1990). The use of
increased mixing with any disinfection technology provides better dispersion of the disinfectant
and forces disinfectant contact with a greater number of microorganisms per unit time. The
increased rate of collisions decreases the required contact time enabling high-rate disinfection
(Glover, 1973). An effective disinfection process will have to provide the desired microbial
deactivation very rapidly under the specific WWF conditions and carry an insignificant amount
of disinfectant residual into the receiving water.
3.2.4 Requirement for Suspended Solids Removal
Effective use of any disinfection technology on WWF requires use of a treatment train,
where its initial segment removes excess suspended solids and its final segment is the
disinfection process. WWF disinfection requires some form of filtration or
clarification/sedimentation prior to introduction of disinfecting chemicals (U.S. EPA, 1973).
High levels of particulate matter in WWF can provide a "shielding effect" in which particles
present in the medium protect the microbes either from disinfecting agent. (Sakamoto and
Cairns, 1997).
Microbial aggregation and particle association are two phenomena that protect
microbes and, thus, are major causes of decreased disinfection efficiency. Microorganisms have
a tendency to clump together to form aggregates. While the organisms living on the outer layer
of the aggregate can be easily disinfected, the microbes living inside are only partially (if at all)
penetrated by the disinfectant or by UV light (Katzenelson et a/., 1976). Particle association can
be represented by attachment of the microorganisms to the particle's surface and by microbial
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occlusion within the particle. Microbes attached to the particle's surface are usually properly
disinfected but microbes occluded or hidden within the particles may not be disinfected at all.
Studies have shown that pretreatment processes (e.g., filtration, sedimentation) can
significantly reduce the effects of both aggregation and occlusion on disinfection efficiency.
Johnson et al. (1983), for instance, tested both filtered and unfiltered secondary wastewater
effluents that were subjected to UV disinfection in side-by-side UV reactors. The filtered
effluent showed significantly better disinfection than the unfiltered medium. The study
concluded that microbial protection by large particle occlusion is the major reason for increased
disinfection efficiency after filtration. Therefore, particle count and size distribution are
important indicators of the influent quality and its need for pretreatment. Particularly sensitive
to suspended solids content is UV disinfection, which is significantly more effective at
suspended solids contents of less than 150 ppm (U.S. EPA, 2002b). UV disinfection tested on
CSO and SSO after compressed media filtration (Fuzzy Filter) showed improved performance
(U.S. EPA, 2002c). In case of chemical disinfection, the lower suspended solids content in the
treated effluent, the less chemical addition and shorter contact times are needed for effective
disinfection.
3.2.5 WWF Disinfection Technologies
Alternatives to chlorination disinfection technologies, for example UV light irradiation,
chlorine dioxide (C1O2), and ozonation (O3), generate significantly less toxic byproducts and
residuals when compared to chlorine (C12). However, only chlorination/dechlorination, as
opposed to alternative technologies, is currently used for WWF disinfection. There have been
several pilot studies on WWF with alternative technologies. The Water Environment Research
Foundation (WERF) were sponsoring a study that evaluates the risks and benefits associated
with various CSO disinfection technologies. A report on its results will be published by 2004.
Disinfection technologies are discussed below in diminishing order of their commercial
availability for WWF treatment.
3.2.5.1 Chlorination and Dechlorination
Disinfection by C12 has proven to be effective, and has been used for wastewater
disinfection in the U.S. since 1855 (White, 1999). Chlorine or its derivatives are the most
commonly applied disinfectants in the U.S. (SAIC, 1998). Chlorine is readily available in
several forms, inexpensive, and effective against bacteria, though not fully effective against
viruses. Chlorine is ineffective in killing protozoa. The easiest way to increase C12 effectiveness
is to increase the C12 dosage within the system. This, however, results in the additional
generation of toxic, carcinogenic, and/or mutagenic byproducts, as well as a high residual
concentration of C12 in the receiving waters. In the last 20 years, disinfection by chlorination has
come under scrutiny. Research studies, particularly for drinking water, have cited health risks
with regard to C12 and its byproducts. Excess of free C12 can cause chlorinated hydrocarbon
formation, i.e., chloroform and trihalomethanes (THMs), which are suspected carcinogens.
Chlorine residuals discharged to natural waters may be harmful to aquatic life.
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Disinfecting high volumes of WWF requires large quantities of C12. Because of the high
risk of gas leaks when transporting gaseous C12, use of liquid C12 in the form of calcium
hypochlorite and sodium hypochlorite is preferred but more expensive. Liquid C12 as sodium or
calcium hypochlorite, is easier to handle, safe to store in onsite tanks, and immediately available
for use. The effectiveness of liquid versus gaseous Cl2for disinfection of WWF has been
investigated. In general, the studies confirmed that liquid C12 is a better disinfectant for WWF,
and WWF facilities are encouraged to changeover from gaseous to liquid C12. When necessary,
excess of free Cl2can be removed by using either gaseous sulfur dioxide or sodium bisulfite
solution. This will eliminate further byproduct formation, but will neither eliminate nor reduce
the already-formed harmful byproducts. Dechlorination also means the addition of another
process, which raises the cost of disinfection. On the average, dechlorination will add about 30%
to the total cost of disinfection. After dechlorination, there is an analytical challenge in
measuring the required low residual level of C12 and the associated monitoring of C12 levels in
receiving waters.
The chlorination/dechlorination pilot study at the 26th Ward WWTP testing facility in
New York City, NY demonstrated that hypochlorite disinfection was a cost-effective technology
for the upgraded Spring Creek facility because of the existing tanks at this facility.
Dechlorination will be added at a later date. Improvements will be made to increase disinfectant
flash mixing and to automate hypochlorite feed and control of the residual chlorine (U.S. EPA,
2002b). The study is described in greater detail in Section 3.2.6.1.
Chlorination/dechlorination of CSO was tested on over 40 wet-weather events at a full-
scale Advanced Demonstration Facility (ADF) in Columbus, Georgia. This study is summarized
under the 3.2.6.3 subsection of this Chapter. Detailed performance results and relative costs are
presented in a report (Columbus Water Works, 2001).
3.2.5.2 Ultraviolet Light Irradiation
Since the early 1900s, UV light irradiation from mercury arcs has been recognized as an
efficient disinfecting agent. At the germicidal wavelengths, within the range of 200 to 320
nanometers (nm) wavelength, UV light disinfects water by altering the genetic material in
microbial cells, preventing reproduction. Peak effectiveness occurs near 253.7 nm, the
wavelength of emission from a mercury arc lamp. UV irradiation has become an acceptable
alternative to chlorination for wastewaters undergoing a secondary or tertiary treatment. Until
recently, it has not been used for low-quality effluents such as WWF as an alternative to
chlorination.
Certain parameters determine the UV dose required for effective disinfection.
Understanding these parameters and their variability in WWF is very important for proper
disinfection system design (Ashok et al. 1997). High variability in WWF flow rates influences
UV disinfection effectiveness, because flow rate is a principal determinant of the dosage of UV
light necessary for effective disinfection (Wojtenko et a/., 200la). This is generally true for all
WWF disinfectants, but UV disinfection is more affected by wastewater quality than chemical
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disinfection technologies. High levels of suspended solids containing particles larger than 2
microns and minerals present in WWF also reduce UV light effectiveness. During disinfection,
the negatively charged quartz sleeves surrounding the UV lamps foul by picking up positive ions
(e.g., Ca, Mg, and Fe) from the water. Fouling materials decrease transmittance of UV light and
thus its disinfection capability (Oliver and Gosgrove, 1975). Use of an in-place cleaning system
can remove fouling materials from the quartz sleeves.
Using UV irradiation for disinfection eliminates many problems arising from
chlorination, such as the need for chemicals and their associated transportation, handling, and
storage, as well as the need for expensive dechlorination facilities. Eliminating large contact
tanks and facility buildings significantly lowers capital and operating costs. UV light irradiation
affects a wide range of microorganisms and does not generate known harmful secondary
chemical byproducts (e.g., THMs). Based on investigations, UV light irradiation for CSO
disinfection shows promise as an effective and safe alternative to chlorination. To inactivate the
target microorganisms efficiently, UV light must penetrate the water. Therefore, the water to be
disinfected must be as clear as possible.
High levels of particulate matter in WWF absorb a large amount of energy, significantly
decreasing the amount of UV light available for disinfection. UV light can disinfect free-living
microorganisms very effectively with a low dose of irradiation, but microbes are often adsorbed
to surfaces of particles (e.g., soil, sediment) or embedded within solid materials (e.g., fecal
material). Solid particles shield the microbes from the disinfecting agent. Adsorption of the
microorganisms to inorganic surfaces does not affect disinfection efficiency as significantly as
adsorption to organic matter. The presence of a surface like clay does not inhibit UV
disinfection because it tends to scatter UV light rather than absorb it.
UV light irradiation is a physical procedure that does not alter the smell or chemical
composition of water. UV disinfection for WWF requires some level of physical pretreatment
(with or without chemicals) to make UV light more effective for WWF (Field, 1996). Pilot
studies have shown that filtration prior to UV disinfection can minimize the effects of particle
occlusion/association (Johnson etal., 1983).
In a 1996 pilot study of high-rate CSO treatment technologies in the Metropolitan
Toronto Area, Canada, UV disinfection was used to achieve an E. coli count of 200 cfu/100 mL
in a CSO effluent treated by a vortex separator, marketed as the Storm King. UV collimated
beam tests were undertaken on only two samples, and in both cases the vortex separator was
operated at a surface load of 10 m/h, with a cationic polymer dosage of 8 mg/L. The residual
total suspended solids (TSS) in vortex effluent samples averaged 42 mg/L and the interim target
fecal coliform count had been achieved at a UV dosage of 30 mWs/cm2, which was considered to
be a feasible dosage for full-scale application. The cationic polymer coagulant was used to
improve the solid/liquid separation efficiency and, thus, facilitate UV disinfection (Marsalek et
al, 1996).
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UV testing on CSOs in the ADF study in Columbus, GA, was also done in a treatment
train arrangement. UV was tested after both vortex and compressed media filtration and its
performance was better on the filtered CSO than on the unfiltered CSO. This study is
summarized under the 3.2.6.3 subsection of this chapter. Detailed performance results and
relative costs are presented in a report (Columbus Water Works, 2001). UV testing after the use
of compressed media filtration (Fuzzy Filter) was done on SSO-type wastewater at the Rockland
County, NY sewer district testing facility. This study is summarized under the 3.2.6.3
subsection of this chapter (U.S. EPA, 2002c).
Investigations of UV light irradiation for CSO disinfection have shown this technology to
have the potential to be an effective and safe alternative to chlorination, assuming the adequate
removal of suspended solids prior to UV application. A CSO disinfection pilot study conducted
at the 26th Ward WWTP testing facility in New York City that evaluated and compared UV light,
O3, C1O2, and chlorination/dechlorination disinfection units showed that the UV light unit was
the simplest unit to operate. This study is summarized under the 3.2.6.3 subsection of this
chapter (U.S. EPA, 2002b).
It is evident from studies and implementation examples described under section 2.6.3,
UV technology has been gaining acceptance for treatment of CSO.
3.2.5.3 Chlorine Dioxide
The use of C1O2 for WWF disinfection has also been investigated as an alternative to
chlorination. The lack of any significant reactions of C1O2 with water is the main reason for its
biocidal effectiveness over a wide pH range. Chlorine dioxide was found to provide excellent
disinfection at a fraction of the C12 dosage, making it cost effective and relatively safe. In
addition to its high effectiveness over a wide pH range, the low reactivity of C1O2 with ammonia
and reduced formation of halogenated organic compounds are its major advantages over C12.
However, the presence of organic and inorganic impurities in water is a limitation of C1O2
disinfection. The impurities create a large oxidation demand for C1O2. These reactions take
place together with disinfection (Katz et a/., 1994). In such a system, the effectiveness of the
disinfecting agent is greatly reduced. Effective treatment of the wastewater by filtration and/or
sedimentation is a precursor for successful C1O2 disinfection (Stinson et a/., 1999). This is of
great importance for CSO applications.
Chlorine dioxide is a very strong and effective wastewater disinfectant. It is not a
chlorinating agent and does not lead directly to the formation of organochlorine compounds
(Dernat and Pouillot, 1992). The major advantages of C1O2 are: its disinfection effectiveness for
Cl2-resistant pathogens (e.g., viruses and protozoa) within a wider pH range, its high solubility in
water, the production of stable and measurable residue, and no reactivity with ammonia to
produce chloramines. Due to these advantages, C1O2 was found to be an attractive candidate for
WWF applications. Because C1O2 is a more powerful disinfectant than C12, lower levels of C1O2
can be used resulting in lower levels of toxic byproducts to get the same level of inactivation.
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For several decades, researchers have compared the respective disinfection efficiencies of
C1O2 and C12. In potable water as well as in wastewater treatment applications, a number of
researchers have found a significantly lower C1O2 demand compared to that of C12. In studies
where equivalent amounts of each of the disinfectants were added to water with various levels of
contamination, after 30 minutes of contact, C12 was found to be largely consumed while C1O2
remained mostly unreacted. This result indicates that C1O2 reacts with fewer compounds than
C12. Due to the limited reactions of C1O2 with organic compounds in water, more of the
disinfectant remained available as a biocidal agent. Chlorine dioxide was found to be a stronger
disinfectant than C12 at shorter contact times and, in addition, was found effective against a
greater number of different microorganisms (Moffa, 1975). Chlorine dioxide was also found to
be a better disinfectant of bacteria and more effective than C12 against viruses and protozoa
(Aieta etal., 1980).
The possibility of using a combination of C1O2 and C12 was investigated for municipal
wastewater treatment by Katz et al. (1994). After adding both agents in equal amounts,
improved disinfection efficiency was observed with all doses, and production of the byproducts,
such as chlorite ion (C1O2) and THMs, was greatly reduced. Chlorine dioxide used in
combination with C12 also resulted in a lower residual C12 concentration. A bench-scale study
was conducted by the U.S. EPA on high-rate disinfection using C12 and C1O2 and its findings
were verified by two full-scale prototype treatment facilities for CSOs (Moffa, 1975). The
concentration of residual C1O2, increased while the concentration of toxic C1O2 decreased. This
was explained by Katz et al. (1994) as being the result of an oxidation reaction between C12 and
C1O2" to produce C1O2. When the combination of C1O2 and C12 is used, C1O2 competes with C12
for the oxidation of organic precursors to THM and chloroorganic compounds. Chlorine reduced
the concentration of C1O2" by oxidizing it back to C1O2. In this case, C12, the cheaper
disinfectant, increased the concentration of C1O2, the more expensive disinfectant, thus lowering
the cost of the disinfection process.
Despite the numerous advantages of C1O2 disinfection, the necessity for C1O2 generation
onsite due to its instability is a major disadvantage. The most commonly used C1O2 generation
method is the reaction of NaClO2 with acid (White, 1999). There are safety considerations
associated with C1O2 generation: instability of C1O2 as a gas, storage and transport of its
precursors (e.g., gaseous C12 sodium chlorite NaClO2) on site, and proper operation of the
equipment. There is a serious problem with a delivery of gaseous C12 as it is prohibited to be
transported through most densely populated areas. There is a new process of C1O2 generation
that uses NaClO2 in the presence of UV light (Stinson et al., 1998). In this process the transport
and handling of gaseous C12 is totally eliminated but this process is still under development and
is not commercially available. Other disadvantages of C1O2 disinfection include lack of data
available for full-scale application to WWF and the potential explosion hazard under certain
conditions.
The potential advantages of using C1O2 as a disinfectant greatly outweigh the possible
disadvantages and inconvenience of onsite generation. When produced, handled, and used
properly, C1O2 is an effective and powerful disinfectant. The sequential addition of C12 with
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C1O2 greatly enhances the disinfection process and is cost effective. Chlorine dioxide appears to
have potential for becoming an effective C12 alternative for WWF disinfection. Further
investigations, however, are recommended to determine its effectiveness in a full-scale WWF
application (U.S. EPA, 2002b).
Chlorine dioxide performed better than chlorination/dechlorination in the Columbus ADF
study (Columbus Water Works, 2001) and in the New York City study (U.S. EPA, 2002b). Of
particular interest was the second phase of the New York City study where a new process of
C1O2 generation using UV light, which avoids the need for gaseous C12, was used as the source
of C1O2. While C1O2 was superior in effectiveness and similar in cost to
chlorination/dechlorination, the UV generation technology for C1O2 needs further development.
Currently, C12 gas cannot be transported within New York City. Thus, because an effective C12-
gas-free process of C1O2 generation has not been proven to be reliable, disinfection with C1O2
cannot be considered for use within New York City, or any other metropolitan area, at this time.
The Columbus ADF study is summarized under the 3.2.6.3 subsection of this chapter. Detailed
performance results and relative costs are presented in a report (Columbus Water Works, 2001).
The New York City study is summarized under the 3.2.6.3 subsection of this chapter. (U.S.
EPA, 2002b).
3.2.5.4 Ozonation
Ozone's ability to inactivate microorganisms was already well known as early as 1886
(White, 1999). It is the strongest and fastest-acting oxidant of all the classical disinfecting
agents used for water sanitation today. Ozone inactivates a wider range of microorganisms than
C12, has a relatively high disinfection-kill power, releases limited byproducts, has the ability to
increase dissolved O2 concentration, is non-reactive with ammonium, and has an excellent ability
for removing undesirable odor and color. In addition to being a strong disinfectant, O3 reacts
with organic impurities (e.g., saturated hydrocarbons, amines, and aromatic compounds)
destroying them in the process and forming such byproducts as acids, aldehydes, bromates,
ketones, and peroxides. Studies evaluating ozonation byproducts are limited, and further
investigation in this area is necessary.
Because O3 is a very strong oxidant, it has the potential for being effective for low-
quality wastewater and WWF disinfection. Organic and inorganic impurities, chemical oxygen
demand (COD), pH, temperature, and suspended solids in waters have a significant impact on O3
disinfection efficiency. The presence of water impurities is a major limiting factor of ozonation
for CSO applications. As a strong oxidant, O3 will react with many organic (e.g., aromatic and
aliphatic compounds, pesticides, humic acids) and inorganic (e.g., sulfide, nitrogen, iron,
manganese, cyanide) compounds producing reaction byproducts (U.S. EPA, 1986a). Reactions
with impurities consume O3, which is then no longer available as a disinfecting agent. As a
result, wastewater with high levels of impurities requires a high dosage of O3 and, thus, an
increased O3 demand, for disinfection to be successful. Although O3 is a strong oxidant and a
powerful disinfectant, its application for WWF disinfection has been very limited. As indicated
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by White (1999), effective ozonation requires relatively good water quality; thus, filtration is
recommended before the ozonation process.
Similar to every other disinfection process, ozonation is most effective for free-floating
organisms. The presence of particles in water makes ozonation challenging. In addition to
particle occlusion, microbial aggregation was also found to be a major factor negatively affecting
ozonation. The rates of inactivation of aggregates were found to be much slower when
compared to free organisms.
The equipment and operating costs associated with ozonation are relatively high. Due to
its high instability, O3 must be produced onsite and used within a short period of time. Skilled
operators and constant attention are required. The necessity for onsite generation makes its
application to the intermittent nature of WWF difficult.
In general, the ozonation process, if properly run, can be successful for disinfection of
various water qualities (wastewater and drinking water). The CSO disinfection pilot study in
New York City showed that there are some safety issues with O3 generation and use, such as
collection of off-gas and destruction of O3, use of water-tight and gas-tight contactors, proper
monitoring of the ventilation system, and use of corrosion-resistant construction materials
(Stinson et a/., 1998; U.S. EPA, 2002b). Ozone instability is also a major factor contributing to
the high cost of this technology. There are currently no WWF facilities using this technology in
the U.S.
In the New York City study, the capital costs of O3 generation were found to be the
highest of all technologies that had been investigated concurrently. Costs of ozone disinfection
were found to be dependent on the cost of electricity as well as the source of oxygen used as a
feed (air vs. pure O2). This study is summarized under the 3.2.6.3 subsection of this chapter.
(U.S. EPA, 2002b).
3.2.6 Description of Disinfection Studies and Implementation Examples
3.2.6.1 Disinfection Pilot Study at the 26th Ward WWTP Testing Facility in
New York City
This pilot study demonstrated alternatives to hypochlorite disinfection for application to
the Spring Creek CSO storage facility and potentially to other CSO facilities. The pilot testing
was divided into two phases. Phase I evaluated treatment performance of five technologies: UV,
O3, C1O2, chlorination/dechlorination, and electron beam irradiation (E-Beam). Based on the
results from Phase I, Phase II provided additional evaluation of technologies that had shown
potential for CSO applications. These were UV, C1O2, and chlorination/dechlorination.
Major findings
With the exception of E-beam, the tested technologies achieved targeted bacterial
reductions of 3 to 4 logs.
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• Chlorination/dechlorination, C1O2, and O3 provided targeted levels of disinfection over
the full range of wastewater quality tested.
• Chlorine dioxide was superior in effectiveness and similar in cost to
chlorination/dechlorination. The new technology for C1O2 generation that does not
require use of chlorine gas needs further development.
The upgraded Spring Creek facility will continue to use sodium hypochlorite for
disinfection, with provisions to add dechlorination at a later date. Improvements will be
made to increase disinfectant flash mixing and to automate hypochlorite feed and residual
control.
Wastewater quality
Five disinfection technologies, UV, C1O2, C12, O3, and E-Beam, were tested for their
effectiveness in reducing bacteria levels in water representative of the CSO at the Spring Creek
storage facility. Tests were conducted during wet and dry events. To achieve a four-log
reduction of fecal coliform and a fecal coliform effluent concentration less than 1,000 colony
forming units/100 mL (cfu/100 mL) required doses for UV, O3, C1O2, and C12 of 60-80
mWs/cm2, 24 mg/L, 8-10 mg/L, and 20-28 mg/L, respectively. The range of disinfectant doses
for each technology reflects the variation in performance between Phase I (December through
March) and Phase II (August through November). The variation in wastewater temperature
between Phase I (mean of 11.6 °C) and Phase II (mean of 20.9 °C) had a significant impact on
the performance of C12 disinfection. The colder winter temperatures impede the formation of
monochloramine, which has approximately 25 times less germicidal efficiency than free C12.
Treatment Performance
Four bacteria indicators were used as a measure of the effectiveness of each of the
disinfection technologies; namely total coliform, fecal coliform, E. coli, and Enterococcus. Kills
of each of the indicators, in terms of log reduction and concentration, were related to dose for
each of the disinfection technologies. Chlorination/dechlorination, C1O2, and O3 at the doses
tested were able to provide the disinfection levels of the four-log reduction over the full range of
wastewater quality tested. UV disinfection effectiveness tended to drop off at higher TSS
concentrations (e.g., TSS greater than approximately 150 mg/L). This was attributed to lower
effective penetration of UV due to harboring of bacteria in solids.
Fecal coliform and E. coli exhibited similar dose-response relationships. However, total
coliform and Enterococcus generally required higher doses to achieve the same level of
inactivation as that for fecal coliform and E. coli. This was observed in all technologies except
for the E-beam, where the inactivation results were inconclusive.
The UV and C1O2 technologies provided nearly complete reductions of bactedophage, a
bacterial virus and microbial indicator. However, the viral inactivation data for the C1O2 system
was limited to only two of the four runs due to operational problems. Of the valid data
considered, the effluent concentrations of bactedophage ranged from non-detect to 60 cfu/mL.
Low influent concentrations of the seeded phage limited the maximum log reduction that could
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be observed. The log reduction of bactedophage ranged from 1.9 to 6.0. Because of the low
concentrations of naturally occurring enteroviruses in the pilot influent, the UV disinfection
could not be evaluated satisfactorily on the basis of the tissue culture infectivity assays,
discussed in Chapter 2. However, based upon the reductions of the marginal concentrations
found and upon the bacteriophage results, these technologies would inactivate most natural
enteroviruses found in wastewater at concentrations on the order of 106 cfu/mL.
UV disinfection achieved 4-log bacteria reduction but at extremely high dosage levels
owing to the impediments of poor water quality. UV effectiveness tended to be reduced by high
TSS concentrations (e.g., greater than 150 mg/L). Additionally, UV effectiveness tended not to
increase at doses greater than 75 mWs/cm2, a phenomena known as "tailing-off"
Ozone disinfection can be accomplished only at high O3 dosage levels. However, the O3
pilot unit did not include a contactor design appropriate for the wastewater conditions tested.
Thus, the required O3 dosages may have been less if a more applicable O3 dissolution/contactor
system were provided. An O3 disinfection system would require contact chambers other than the
tankage that presently exists at Spring Creek.
Chlorine disinfection included dechlorination to eliminate residual C12. Chlorination as
well as dechlorination can be accomplished using the existing tanks at the Spring Creek
Advanced Wastewater Pollution Control Plant (AWPCP). High-rate mixing can be added to the
head end of the tanks. Chlorine dioxide disinfection can be accomplished at doses on the order
of 30% of the required C12 dose.
Chlorination/dechlorination and C1O2 were determined to be the most cost effective
technologies for application to Spring Creek. However, neither of the C1O2 generation methods
tested are currently feasible for use within New York City. The C12 gas/solid sodium chlorite
generation method is not feasible because of its use of C12 gas, and the UV/sodium chlorite
generation method is not feasible because of its developmental status as a prototype. The capital
costs for UV and O3 were significantly more expensive than chlorination/dechlorination or C1O2.
For other CSO facilities that do not have existing tanks for contact time, UV could be
economically attractive.
In the case of C1O2, there is no significant increase in disinfection performance beyond a
contact time of three minutes. This is in contrast to the chlorination results, which show a greater
dependence on contact time and required five minutes for comparable kills. The difference is
attributed to ClO2's greater bactericidal properties and solids penetration characteristics than
those of chlorination. The results of this study confirm the optimum contact times for C1O2 and
chlorination/dechlorination of three and five minutes, respectively, originally determined in the
Syracuse and Rochester studies (U.S. EPA, 1979a and 1979b). Chlorination/dechlorination and
C1O2 were determined to be the most cost-effective technologies for application at this facility.
Further development of the UV-chlorite C1O2 generator is required before reliable costs for this
technology can be developed.
3-15
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Disinfection Residuals and Toxicity
Selected disinfection effluent residuals and byproducts, namely C1O2, chlorate, chlorite,
total residual chlorine (TRC), volatile and semivolatile organics, haloacetic acids, were
monitored to relate these residuals to disinfectant dose. UV disinfection had the distinct
advantage of producing no byproducts. This is in contrast to C12 and C1O2, which produced
increased levels of TRC, chlorate, chlorite and haloacetic acids in the effluent. The slightly
increased haloacetic acid concentrations were considered insignificant. The increased TRC,
chlorate and chlorite concentrations were directly related to increased C12 and C1O2 dose.
No additional toxicity was observed in the UV effluent as compared to the UV pilot
influent. In contrast, there were occurrences where the C1O2 effluent was considerably more
toxic than the pilot influent. An attempt was made to correlate this toxicity with the specific
disinfection byproducts, in particular TRC, chlorate and chlorite, but no correlation could be
made. It is likely that the increased effluent toxicity is directly related to influent toxicity (i.e.,
influent water quality) or a synergistic effect of the disinfectant residuals, which could not be
measured. Although the concentrations of TRC, chlorate and chlorite did not cause concern
about effluent toxicity, this relationship should be revisited when establishing C1O2 dose for
specific sites.
Effluent TRC was generally below 0.1 mg/L following dechlorination as compared to a
receiving water quality standard of 0.0075 mg/L. This TRC value of dechlorinated effluent
reflects the practical quantitation limit of the process instrumentation used. Lower TRC values
could not be quantified. Often, the dechlorinated effluent TRC instrumentation displayed a
negative value indicating the presence of excess bisulfite. Residual C12 was also monitored in
the C1O2 effluent. However, these TRC values include all oxidizing species of C12 and the
possible presence of free and combined C12 could not be differentiated from C1O2, C1O2" and
cio3-.
Chlorine Dioxide Generation
The method of generating C1O2 must be considered when selecting the appropriate
disinfection process. The C12 gas/solid sodium chlorite generation method was tested during the
Phase I and Phase II pilot studies. Although this pilot unit was reliable, the use of C12 gas (either
with C12 cylinders or with on-site C12 gas generation) in this process may limit its application in
residential and urban areas, including New York City. The UV-sodium chlorite solution
generation method was also tested during the Phase II pilot study. This method had the distinct
advantage of not using or generating chlorine gas in the generation process. However, this
technology is currently in the prototype stage of development and would need to be developed as
a full-scale unit to be considered further. The UV-chlorite generator from the UVD, Inc., was a
prototype unit.
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Cost Comparison
During the Phase I pilot study, conceptual level cost projections were prepared for each
disinfection technology for comparison purposes, with the goal of recommending a technology
for implementation at the Spring Creek storage facility. The Phase II pilot study results served
to verify the Phase I result; as such, the assumptions and approach used for the original cost
comparison were applicable. Costs for each disinfection technology were prepared on a
common flow basis and were prepared for a range of flow rates experienced at Spring Creek.
See Table 3-2. This approach shows the sensitivity of cost to flow rate, and allows independent
comparison of technology costs at similar flow rates. Equipment capital costs were developed
for peak design flow conditions of 1,250 cubic feet per second (cfs) (800 million gallons per day
(mgd)), 2,500 cfs (1,600 mgd), and 5,000 cfs (3,200 mgd) for a duration of 4 hours. (U.S. EPA,
2002b).
3.2.6.2 Continuous Deflection Separation, Fuzzy Filter and UV Treatment of
SSO-Type Wastewaters: Pilot Study Results
This study evaluated three technologies for treatment of SSO and CSO overflows. These
were the Continuous Deflection Separation (CDS) and Fuzzy Filter high-rate solids removal
technologies, and UV high-rate disinfection. The study was conducted at the Rockland County
Sewer District No.l, in Orangeburg, NY from August 1998 to January 2001.
Three different lamp systems were evaluated within the UV disinfection studies. These
were:
• PCI Wedeco UV Technology (now Wedeco Ideal Horizons). This system represents
newer low-pressure lamp UV systems, which takes advantage of the high power
conversion efficiency of the low-pressure lamps, while getting higher UV outputs.
• Aquionics UV Technology. This system utilizes medium-pressure lamps. These are less
efficient than conventional lamps but their total UV output is higher resulting in a lower
number of lamps to achieve the required light intensity.
• Generic Medium-Pressure, Open-Channel System. The channel was designed to operate
lamps at two different spacings: 4- and 6-inch.
The overall objective of the study was to evaluate high-rate solids removal technologies
on SSO and CSO type wastewaters, and the subsequent UV disinfection of the treated
wastewaters. The results are given below.
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Table 3-2. Cost Projection of Disinfection to be Implemented at the Spring Creek Facility
Peak Design
Flow (cfs)
Capital Costs
Annualized
Capital Costs
Annual O&M
Cost
Total
Annualized
Costs
Conceptual Level Facility Disinfection Costs ($)
Chlorination/Dechlori nation
1,250
912,000
93,000
255,000
348,000
2,500
1,045,000
107,000
255,000
362,000
5,000
1,219,000
124,000
255,000
379,000
Chlorine Dioxide
1,250
695,000
70,000
294,000
364,000
2,500
1,159,000
119,000
294,000
413,000
5,000
1 ,932,000
196,000
294,000
490,000
Ozone
1,250
19,221,000
1,957,000
534,000
2,491,000
2,500
24560,000
2,502,000
587,000
3,089,000
5,000
30,539,000
3,111,000
657,000
3,768,000
uv
1,250
48,052,000
4,894,000
248,000
5,142,000
2,500
67,272,000
6,852,000
497,000
7,349,000
5,000
87,774,000
9,592,000
992,000
10,584,000
Notes: 1. Costs are present worth in 2000 dollars.
2. Capital costs are based upon sizing to meet peak design flow and a 4-log reduction in fecal coliform.
3. Capital costs are for installation of Spring Creek and are for process equipment only. Costs do not
include additional contact tankage (if required) or support facilities.
4. Annual operating costs are based upon an assumed typical 40 CSO events/year at a volume treated
of 15 million gallons per event.
5. Annualized costs are based upon a period of 20 years at an interest rate of 8%.
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UV Disinfection Dose Requirements and Particle Size Impacts
The dose-response analyses indicated that removal of particles greater than 50-micron in
size will improve the efficiency of the UV process because filtration to such levels removes a
substantial amount of occluded bacteria. Samples were blended prior to analysis to release
occluded bacteria so they could be detected in analysis. Blending the unfiltered samples released
fecal coliform and improved recovery of occluded bacteria. Blending samples that had been
filtered at retention levels between 1 and 50 microns did not have a significant impact on
coliform recovery and did not impact UV dose requirements to accomplish targeted reductions.
The UV dose requirement to accomplish 3-log reduction of fecal coliform in primary-
type wastewater (i.e., waste water of a quality equivalent to a primary-treated waste water),
pretreated to remove particles greater than 50-microns is approximately 20 mJ/cm2. The results
suggest that the maximum reductions that can be expected under practical dose applications up
to 40 mJ/cm2 are 3.5 to 4 logs. With unfiltered effluents and primary-treated wastewaters passed
only through the CDS unit, the maximum reductions suggested by the dose-response analyses
are approximately 2.5 to 3.0 logs (based on enumeration of blended samples).
CDS Process Performance
The CDS process is capable of accomplishing approximately ten percent TSS removals
with a 1200-micron screen. This increases to approximately 30 percent with a 600-micron
screen. In both cases, it appears that removals were independent of the flow rate, within the
range of flows tested.
The CDS unit, based on visual observations, was effective in capturing and removing
debris, including paper and plastics, fibers, and preventing transport to downstream processes.
In this respect, the wider aperture screens were as effective as the smaller aperture screens and
are more easily maintained. The wider aperture screen tended to be self-cleaning while the
smaller aperture screen required manual cleaning and tended to retain the debris on the screen
surface. The CDS process can provide protection of downstream filters or other pretreatment
devices by removing debris and floatables.
Fuzzy Filter Performance
The filter was effective in removing larger-size suspended solids. The particle size
distribution (PSD) and dose-response analyses confirmed that these removals centered on
particles greater than 50 microns. The system is more effective in this application at 20-percent
compression and at hydraulic loadings between 400 and 800 Lpm/m2 (10 and 20 gpm/ft2). At
these conditions, TSS removals averaged approximately 40 %. Removals were consistently less
at these hydraulic loadings for the 10 and 30 % compressions.
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UV Disinfection Performance
The combined results generated with the three UV units indicate that a degree of
disinfection with primary wastewaters can be accomplished by UV radiation. Reductions
between 2.3 and 2.8 logs can be achieved at hydraulic loadings between 8 and 38 Lpm/kW of
lamp input power (2 and 10 gpm/kW) based on the enumeration of blended samples. This is
equivalent to approximately 3 to 3.5 logs when enumeration is conducted using standard
analyses without blending samples. Doses greater than 40 mJ/cm2 are required to achieve these
reduction levels (U.S. EPA, 2002c).
3.2.6.3 Advanced Demonstration Facility (ADF) in Columbus, GA
Chlorination/dechlorination of CSO, along with several alternative technologies, were
tested on over 40 wet-weather events at a full-scale ADF in Columbus, GA. The CSO testing
program at ADF was a part of a multi-year watershed study sponsored by the Columbus Water
Works Agency with the Wet Weather Engineering & Technology (WWETCO) firm as the
principal contractor and with the involvement of the WERF and the U.S. EPA. ADF is
comprised of multiple CSO technologies arranged as treatment trains: hydraulic controls,
screening, vortex separation, compressed media filtration, and chemical disinfection using C12 as
sodium hypochlorite, C1O2, peracetic acid, and UV disinfection. Multiple technologies were
tested side-by-side and in sequential and split stream for determining performance at different
loading rates and equipment settings. Performance results and relative costs are summarized
below (Columbus Water Works, 2001).
ADF CSO Technology Evaluations
The ADF demonstration facility, with permitted capacity of 48 MGD, consists of coarse
screening and flow controls, six 32-ft diameter vortex separators, a compressed media Fuzzy
Filter (a 30-inch bed of 1-inch fiber balls contained between two perforated plates), a medium
pressure UV system located downstream of the Fuzzy Filter (u-tube arrangement of two banks of
42 bulbs), and other auxiliary equipment. The ADF is fully automated and operates during wet-
weather events when runoff exceeds interception. Manned operations include both pre-and post-
event activities as well as preventive maintenance. Continuous rainfall monitoring and level
instruments automatically initiate operations such as screening, underflow pumps, and
disinfection equipment. Post-event activities include residuals removal from screens and grit
bins, sodium bisulfite dechlorination, and equipment operation checks.
Testing of three chemical disinfection technologies, C12 as sodium hypochlorite, C1O2,
and peracetic acid, was conducted in designated vortex separators for each technology. The
vortex separator is designed to remove grit and concentrated solids but can be and was used for
combined solids removal and chemical disinfection. Vortex has no moving parts and acts like a
plug-flow reactor providing contact time greater than 70% of theoretical. There is higher usage
of chemicals in a vortex than in a separate disinfection tank but the cost of additional chemicals
is less than the cost of separate tankage. Sodium bisulfite dechlorination was also conducted in a
vortex.
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Chemical disinfectant was delivered by feed pumps according to a developed control
algorithm for changeable dosing. At the ADF, the disinfectant demand for CSO was correlated
with its ammonia and COD content in conjunction with the continuous flow and time
measurements. Chemical disinfection efficiency also correlated with pH, temperature, and TSS.
The highest disinfectant dose was given at the beginning of the event and it was diminishing as
the event was progressing. A minimum contact time used was three minutes. Disinfectants
listed in order of their effectiveness were C1O2, sodium hypochlorite, and peracetic acid,
however all were capable to accomplish a satisfactory disinfection. Chemical dosing under
similar conditions requires 15 mg/L sodium hypochlorite, 16 mg/L peracetic acid, and 12 mgL of
C1O2.
Sodium hypochlorite was selected because C1O2 requires generation onsite with the use
of C12 gas and peracetic acid is not licensed in the U.S. for wastewater disinfection. Sodium
hypochlorite (C12) dose varied from 4 to 30 mg/L with average concentrations between 8 and 9
mgL. Contact times ranged from 6 to 40 minutes at peak flow rates at events tested. The
predominant contact times were between 10 and 20 minutes. Chlorine disinfectant residuals,
when operating with variable feed rates, were typically around 1 mg/L. Dechlorination was
designed for chlorine residuals exceeding 1 mg/L.
The compressed media filter provided a sufficient pretreatment level for UV disinfection.
A double bank of medium pressure high intensity UV lamps (42 lamps per bank) reduced
bacteria counts to the hundreds and thousands level (colonies per 100 mL) for flows of 10 to 20
MGD. The contact time for UV disinfection was two minutes. These results were for average
conditions of TSS at 50 mg/L, 20% light transmittance and 25 degrees Centigrade water
temperature. Transmissivity of treated flow was very important for UV. For example, UV
disinfection of E. coli bacteria in filtered effluent with about 60% transmissivity was on the
order of a magnitude higher (in hundreds of colonies per 100 mL) than in effluent with 40%
transmissivity (in thousands of colonies per 100 mL). In contrast, the unfiltered CSO UV
transmittance was as low as 20%.
A spreadsheet model was developed to evaluate combinations of intercept, storage, and
flow through CSO treatment processes. The evaluation considered removal efficiencies, capital,
and operational costs. The ADF findings provided performance criteria for vortex separation,
C12, C1O2, and peracetic acid disinfection, and compressed media filtration followed by UV
disinfection.
An optimized model of the ADF facility was developed. The optimized facility includes
two 32-ft diameter vortex separators, instead of current six vortex separators, with C12
disinfection followed by dechlorination and 2,000 cubic feet of compressed media filtration,
instead of the current 1,000 cubic feet, followed by UV disinfection. The intercept capacity in
this example is 10 MGD. The recommended peak flow capacity of the facility is 90 MGD.
Present worth, capital and operation and maintenance (O&M) costs were developed for
various treatment trains, including the optimized facility, using 1995 construction costs and
annual O&M costs based on several years of operation. Capital costs for a treatment system
3-21
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designed for 63% removal of TSS were estimated to be approximately $10,000 per acre of
combined sewer service area; annual operating costs were estimated to be about $163 per acre.
Designing the system for 80% removal of TSS increased the capital cost nearly threefold, with
annual operating cost doubling. As discussed above, removal of TSS is representative of
disinfection effectiveness, especially for UV (Arnett, 2003. Personal Communication).
3.2.6.4 Washington, DC. Northeast Boundary Swirl Facility (NEBSF)
(Disinfection Implementation)
The NEBSF, operated by the District of Columbia Water and Sewer Authority (WASA),
provides treatment and disinfection for up to 400 MGD of CSO before discharging to the
Anacostia River. The facility provides mechanical screening followed by three 57-ft diameter
swirl concentrators. The effluent from swirl concentrators flows to a mixing chamber where
sodium hypochlorite is added, usually at a dose of 5 mg/L. Sodium bisulfite is added at the end
of the outfall for dechlorination, usually at a dose of 2 mg/L. Flows above 400 MGD are
discharged untreated. Samples taken during CSO events at the mixing chamber and at the river
outfall are analyzed for Enterococcus and fecal coliform. Reported counts range from less than
10 MPN/100 mL to in excess of 250,000 MPN/100 mL. The high numbers are associated with
events in excess of 400 MGD and represent blending of treated and untreated CSO.
Annual operating costs for the NEBSF are estimated to about $230,000. This is based on
$180,000 for labor and $50,000 for chemicals. The facility discharges on average about 100
times per year, with an average total volume of approximately 1,500 MG (Siddique, 2003.
Personal Communication).
3.2.6.5 Birmingham, AL. UV Disinfection at Peak Flow WWTP
(Disinfection Implementation under Construction)
The Jefferson County Environmental Services Division for the City of Birmingham and
about 20 neighboring communities is in the process of constructing a 350 MGD peak excess
flow treatment facility. The new facility, named the Village Creek Peak Flow Wastewater
Treatment Plant (PFWWTP), includes a pump station, with 360 MGD capacity, 20 surge basins
with surface aeration for mixing (total capacity of 90 MG), granular, monomedia, deep bed
filters with 350 MGD capacity, UV disinfection, and a 24 megawatt generating facility
(primarily to power the pump station and UV). The UV system will have a total of 2,688 lamps
at a peak power requirement of 7,526 kW. The total installation cost of the UV facility is
estimated to be $13 million; the cost of UV equipment is about $10.7 million. Operating costs
are not available (Chandler, 2003. Personal communication).
3.2.6.6 Oakland County, ML Chlorine Disinfection at Acacia Park
(Disinfection Implementation)
The Office of the Oakland County Drainage Commissioner currently operates three CSO
retention basins in southeastern Michigan, all of which provide treatment and disinfection of
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flows that exceed their storage capacity. The Acacia Park CSO Retention Treatment Basin
(RTB) is a 4 MGD basin that serves a combined area of about 818 acres. Disinfection is by
sodium hypochlorite. The feed system provides a dose of 10 mg/L at a CSO flow rate of 426
MGD. There is no dechlorination. The disinfection target is a fecal coliform count of less than
400 cfu/100 mL at a total residual chlorine level of 1.0 mg/L.
Annual operating costs for the Acacia Park facility are estimated to be $120,000. This
includes $58,000 for labor, $24,000 for energy and utilities, $26,000 for chemicals, and $10,500
for laboratory and other services. Over the three-year demonstration period, the facility captured
60% of the flow it received; that is treated overflows represent 40% of flow into the facility. The
total volume of flow into the facility was estimated at 146 MG, with 88 MG retained and
returned to the sewer system and 58 MG treated and discharged. Overflows occurred on average
four to five times per year, and ranged in volume from 0.13 to 17 MG (Mitchell, 2003. Personal
Communication).
3.2.6.7 Bremerton, WA. UV Disinfection at CSO Treatment Facility
(Disinfection Implementation)
The City of Bremerton has recently constructed a CSO treatment facility that uses high-
rate clarification, followed by UV disinfection, to treat flows up to 45 MGD. The facility uses a
medium-pressure, high-intensity UV system that employs a total of 90 lamps. A 500-kW
generator supplies power to the UV system as well as pumps, mixers, and other equipment. The
clarification system uses a polyaluminum chloride coagulant. The primary reason for choosing
UV over chlorination was to avoid degradation of hypochlorite between discharge events, which
occur about 20 times a year. Bremerton installed a UV system at a a cost of about $600,000 to
disinfect CSO discharges. The annual operation cost for the entire facility is estimated to be
about $50,000 (Poppe, 2003. Personal Communication).
3.2.6.8 Disinfection of Collected Stormwater and Dry Weather Urban Runoff
New Orleans, LA - Stormwater Disinfection
The city of New Orleans, LA evaluated a prototype disinfection facility for Stormwater
using sodium hypochlorite in the late 1960s and early 1970s; (Pavia and Powell, 1968) however,
they did not adopt the practice permanently.
Santa Monica Urban Runoff Recycling Facility (SMURRF)
Santa Monica's urban runoff recycling facility (SMURRF) project, completed in
December 2000, in Santa Monica, CA, treats dry weather runoff water from excessive irrigation,
spills, construction sites, pool draining, car washing, the washing down of paved areas, and some
wet weather runoff. SMURRF treats an average of 0.5 MGD of the above urban runoff with
solids, and oil and grease removing technologies prior to UV disinfection for removal of
pathogens. The treated runoff is reused for landscape irrigation and for in dual-plumbed
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buildings for flushing of toilets. For more information, see the Internet site at:
http://Epwm.Santa-Monica.Org/Epwm/Smurrf.html.
Moonlight Beach Urban Runoff Treatment Facility
Moonlight Beach Urban Runoff Treatment Facility in the City of Encinitas, CA has been
treating dry season urban runoff since September 2002. The facility accepts flows up to 150
gpm. The technologies used are filtration followed by UV disinfection. Coliform bacteria were
reduced by over 99%. The facility does not operate during the wet season (Rasmus and Weldon,
2003).
3.3 Best Management Practices (BMPs) for Control of Pathogens in
Urban Stormwater
3.3.1 Introduction
Practices to control and manage the quality and quantity or urban runoff have become
widespread. This set of practices has been labeled best management practices or BMPs.
Structural BMPs are designed to function without human intervention at the time a storm event
occurs (Urbonas, 1999). Wet ponds, dry ponds, constructed wetlands, filters, rooftop storage,
and swales are examples of structural BMPs that can be applied to urban stormwater.
Eliminating illicit cross connections between the sanitary sewage system and separate
stormwater drainage system is another structural BMP. Similarly, reduction of stormwater
volume that enters combined or sanitary sewer systems aids in reducing CSO and SSO volumes.
These measures are distinct from the others because they involve repairing the stormwater or
sewerage system, rather than erecting a structure to manage or control stormwater quality. Other
practices that reduce stormwater volume known as inflow reduction techniques include
disconnection of roof leaders and redirection of area and foundation drains and basement sump
pumps to soils where the flow will infiltrate to the ground or groundwater. Nonstructural BMPs
are generally good housekeeping practices or measures designed to institute good housekeeping
for reducing or preventing pollutant deposition in a watershed, e.g., public education or
regulation (Urbonas, 1999).
This section provides a detailed discussion of the application of structural and
nonstructural BMPs to stormwater microbial contamination. Available data on performance of
BMPs for removing microorganisms from stormwater are presented. However, quantitative
results are inconclusive or unavailable for many of the BMPs.
3.3.2 Structural BMPs
Wet ponds, dry ponds, constructed wetlands, filters, rooftop storage, and swales exhibit
varied effectiveness for volume reduction and removal of suspended solids, metals, and
nutrients. Structural BMPs have been applied to control pathogens to a lesser extent than to the
other pollutants, and have produced mixed results. Often, controlling pathogens or
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microorganisms is a secondary goal for these BMPs, which are more routinely implemented for
reducing flow volume, sediment, or nutrients. Some environmental professionals are of the
opinion that these practices do not affect pathogens to a meaningful degree and, therefore,
should not be implemented to obtain the goal of reducing microbial concentrations.
Microorganism or pathogen removal has been reported most frequently by sand filters, wetlands,
and wet detention ponds. EPA Storm Water Technology Fact Sheets for these BMP types are
available on EPA's web site at http://cfpub.epa.gov/npdes/ (U.S. EPA, 2003b). The fact sheets
include the following information:
• description
• applicability
• advantages and disadvantages
• design criteria
• performance
• operation and maintenance
• costs
Limited research has been conducted on the effectiveness of structural BMPs for
controlling stormwater pathogen loads to receiving waters. Much of the existing information has
been compiled by Winer (2000) and by ASCE (2002) in U.S. EPA-sponsored projects. The data
is compiled in database format, therefore, it is general in nature. It is included here to provide
the reader with the range of BMP effectiveness and the database reference information. For
more detailed information on a particular site, the reader should go to the original reference cited
in the database. Reported fecal coliform removal efficiencies range from 99% at a wet pond in
Ontario, Canada to -134% in a Fremont, CA wetland. These data show that while there are
cases where microorganism reduction can be achieved to some extent by employing BMPs,
BMPs also serve as environments where microorganisms are generated, presumably from
increased wildlife populations and resuspension of bottom deposits. Table 3-3 presents
performance data on the effectiveness of four types of BMPs for treating stormwater: wetlands,
dry ponds, wet ponds, and sand filters (ASCE, 2002; Kurz, 1998; Winer, 2000). Figure 3-1
illustrates the variability of fecal coliform percent removal efficiencies reported. For each case
study, the removal efficiencies are calculated using the average inlet and outlet fecal coliform
concentrations.
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Table 3-3. Stormwater BMP Effectiveness Data.
BMP
Type
Wetlands
Dry Pond
Sand
Filter
Total Coliform
(CFU/100mL)
Influent
3
Effluent
2120
% Removal
-706
59.4
Fecal Coliform
(CFU/100mL)
Influent
2516
2516
2
690
1350
5695
Effluent
5882
4581
236
20
768
18528
% Removal
78
-134
-82
-117
97
55
78
37
83
36
37
-85
81
66
Location and Reference
Lake Beardall, FL. Submerged gravel
wetland. Egan etal., 1995, in Winer,
2000 (Study 91).
Fremont, CA. ASCE, 2002.
Fremont, CA. ASCE, 2002.
Sea Pines Plantation, SC. Surface
flow, full scale, natural marsh,
abundant wildlife, runoff and manure
from horse trail. MacClellan, 1989,
referenced in Table 17-3 of Kadlec
and Knight, 1996.
Kingston, MA. Shallow marsh (natural
or constructed not specified).
Horsley, 1995, in Winer, 2000 (Study
79).
Glenwood, WA. Shallow marsh
(natural or constructed not specified).
Koon, 1995, in Winer, 2000 (Study
80).
Maple Run III, TX. ASCE, 2002.
Joleyville, TX. City of Austin, Texas,
1990, in Winer, 2000 (Study 105).
Brodie Oaks, TX. City of Austin,
Texas, 1990, in Winer, 2000 (Study
106).
Barton Creek, TX. City of Austin,
Texas, 1990, in Winer, 2000 (Study
107).
Highwood, TX. City of Austin, Texas,
1990, in Winer, 2000 (Study 108).
Barton Ridge Plaza, TX. City of
Austin, Texas, 1990, in Winer, 2000
(Study 109).
Barton Creek Square, TX. City of
Austin, Texas, 1991 , in Winer, 2000
(Study 110).
Madeira Beach, FL. Kurz, 1998.
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Table 3-3 continued. Stormwater BMP Effectiveness Data.
BMP
Type
Wet
Ponds
Total Coliform
(CFU/100mL)
Influent
470
Effluent
395.6
% Removal
16
64
Fecal Coliform
(CFU/100mL)
Influent
83633
6937
17619
Effluent
1324
1779
2516
4764
783
% Removal
70
98
86
90
64
73
-6
46
64
56
99
97
98
Location and Reference
Monroe Street, Wisconsin.
Bannerman and Dodds, 1992, in
Winer, 2000 (Study 91).
St. Elmo, TX. City of Austin, Texas,
1996, in Winer, 2000 (Study 26).
Unqua, NY. Driscoll, 1983, in Winer,
2000 (Study 34).
Heritage Park, Ontario, Canada.
Liang, 1996, in Winer, 2000 (Study
43).
Jacksonville, FL. ASCE, 2002.
Fremont, CA. ASCE, 2002.
Davis, NC. FC Mass Removal
Efficiency reported 48.1%. Borden et
a/., 1996, in Winer, 2000 (Study 11).
Piedmont, NC. Borden et a/., 1996, in
Winer, 2000 (Study 12).
Woodhollow, TX. City of Austin,
Texas, 1991, in Winer, 2000 (Study
13) and ASCE, 2002.
Harding Park, Ontario, Canada.
Fellows et a/., 1999, in Winer, 2000
(Study 16).
East Barrhaven, Ontario, Canada.
Ontario Ministry of the Environment,
1 991 , in Winer, 2000 (Study 1 9).
Kennedy-Burnett, Ontario, Canada.
Ontario Ministry of the Environment,
1 991 , in Winer, 2000 (Study 20).
Uplands, Ontario, Canada. Ontario,
Canada. Ontario Ministry of the
Environment, 1991, in Winer, 2000
(Study 21).
Tampa, FL. Kurz, 1998.
Influent and effluent data provided in table when available.
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Figure 3-1. Fecal Coliform % Removal Efficiency by BMP Type.
150
(D
'O
it
LJJ
"(0
O
0)
o:
100 -
50 -
0 -
-50 -
-100 -
-150
n=14
n=6
n=1
n=7
78.0
70.8
»
-17.2
36.4
Wetlands
Dry Ponds Wet Ponds Sand Filter
Legend: • is mean; error bar is standard deviation.
ASCE, 2002; Kurz, 1998; Winer, 2000.
BMP Type
n = number of BMPs reported
There are many factors affecting variability including stormwater characteristics, BMP design,
and environmental factors contributing to microorganism die-off.
3.3.2.1 Ponds and Wetlands
In contrast with the fact that better performance was observed in wet ponds over wetlands
in the studies reviewed above, a number of research studies show that wetlands may provide
advantages over ponds for indicator microorganism removals. One study found greater removal
of thermotolerant coliforms, enterococci, and heterotrophic bacteria from stormwater in a
wetland system (80-87%) than in a pond (-2-22%) (Davies and Bavor, 2000). The researchers
attribute greater bacteria removal in the wetland to increased sedimentation aided by vegetation
and increased removal of fine suspended particles (< 2 microns) with attached bacteria. Pond
and wetland performance on microorganisms in sewage is an indicator of their performance on
stormwater. A wastewater treatment wetland removed 97-99.9% of fecal coliform and
Enterococcus and 70% of coliphage (Stenstroem and Carlander, 2000). The investigators
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attribute the bacteria concentration reductions to the wetland's ability to remove suspended
particles. Viruses have been shown to accumulate in wetland biofilms resulting in their removal
from the effluent (Flood and Ashbolt, 2000).
The University of Arizona sponsors a research program on constructed wetlands
treatment of secondary sewage effluent at the Pima County Constructed Ecosystem Research
Facility in Tuscon. Although the research examines the effect of constructed wetlands on
reducing microbial pathogen and indicator concentrations in secondary sewage effluent, the
results provide useful information that can be applied to stormwater. A duckweed-covered pond,
a multi-species subsurface flow wetland, and a multi-species surface flow wetland reduced
concentrations of Giardia cysts, Cryptosporidium oocysts, total coliform, fecal coliform,
coliphage, and enteric viruses in secondary sewage effluent (Gerba et al., 1999; Karpiscak et al.,
1996; Thurston et al, 2001). Removal of the larger microorganisms, i.e., Giardia and
Cryptosporidium, was the greatest in the duckweed pond, with sedimentation thought to be the
primary removal mechanism. In contrast, the greatest removal of coliforms and coliphage
occurred in the subsurface flow wetland, which may be related to the large surface area available
for adsorption and filtration (Gerba et al., 1999). When supplying potable water to a wetland at
the facility, Thurston et al. (2001) showed that total and fecal coliform concentrations increased.
The researchers attribute the greater densities found in the summer months to the flora and fauna
in and around the wetland. Warm waters promote the growth of bacteria contained in the animal
feces deposited in the wetland. Increased plant growth may increase root exudates, oxygen to
the rhizosphere, and accumulation of organic matter, believed to increase microorganism growth
(Thurston et al., 2001). The results of these studies are summarized in Table 3-4.
Performance of constructed wetlands treating dairy farm wastewater for use in irrigation
provides another source of information related to the effectiveness of constructed wetlands on
removing pathogens from stormwater runoff. Kern et al. (2000) conducted a seasonal effects
study at a side-by-side wetland constructed at the Institute of Agricultural Engineering in
Potsdam, Germany. The subsurface flow wetland with a horizontal water flow reduced fecal
coliform densities by 99.3 and 95.8% in the summer and winter, respectively. The principal
mechanism in eliminating fecal coliform seemed to be adsorption to soil particles followed by
die-off and predation (Kern et al., 2000). During the summer months, vertical distribution of
fecal coliform densities in the control wetland bed, which did not receive wastewater, was
equivalent to the levels in the treatment bed. In the winter, fecal coliform counts were three
orders of magnitude higher in the treatment bed. High counts in the control bed in the summer
were attributed to the presence of warm-blooded animals.
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Table 3-4. Results of Wetlands Effectiveness Studies on Secondary Sewage Effluent at Pima
County, AZ Constructed Ecosystem Research Facility.
Reference
Karpiscak et al.
(1996)
Gerba et al.
(1999)
Gerba et al.
(1999);
Thurston et al.
(2001)
Wetland Type
Multi-Species
Surface Flow
Duckweed
Covered Pond
Multi-Species
Subsurface
Flow
Percent Reduction
TC
98
62
99
FC
93
61
98
Giardia
73
98
88
Cryptosporidium
58
89
64
Enteric
Viruses
98
38*
95
Coliphage
N/A
40
N/A
* from Karpiscak et al. (1996) reporting July - December 1994; other duckweed results reported by
Gerba et al. (1999) for period July 1994 - December 1995
Karpiscak et al. (1999) studied the effectiveness of an integrated wastewater treatment
facility, consisting of solids separators, anaerobic lagoons, aerobic ponds and constructed
wetlands, on dairy waste in Glendale, Arizona. In the aerobic pond, fecal coliform and Listeria
concentrations decreased by 98.5 and 96.6%, respectively. Total coliform, however, increased
by approximately 40%. Concentrations of all three organisms were decreased in the wetlands,
total coliform by 79%, fecal coliform by 82.8%, and Listeria by 99.1%. Reductions are
attributed to UV radiation, degradation of organic matter, solids settling, competition from other
microorganisms, phytoremediation, and residence time.
3.3.2.2 Sand Filters
Sand filters operate by trapping suspended particles or adsorbing pollutants. Sand filters
can be constructed in underground trenches or in above-ground, pre-cast concrete boxes.
Advantages include the lower areal requirements than ponds and the ability to install them out of
public view (Kurz, 1998), both of which facilitate their use in ultra-urban environments where
ponds are more difficult to site.
3.3.2.3 Illicit Discharge Detection and Elimination
Improper connections to storm drainage systems convey contamination to receiving-
water bodies. Sources of microbial contamination transported through this route include sanitary
wastewater and septic tank effluent (Pitt et al, 1993). Since the 1980s, many municipalities
initiated programs to identify and correct illicit connections in response to information
highlighted by EPA's Nationwide Urban Runoff Program (U.S. EPA, 1983) and the 1987 Clean
Water Act. The Clean Water Act requires municipal separate storm sewer system discharge
permits to effectively prohibit non-storm water discharges into storm drains. EPA has an Internet
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site that presents information about illicit discharges, how specific municipalities are working to
address them, and methods for identifying them:
http://cfpub2.epa.gov/npdes/stormwater/menuofbmps/illi 2.cfm (U.S. EPA, 2003b). Pitt et al.
(1993) published an EPA User's Guide on investigating inappropriate pollutant entries into
storm drainage systems available at http://www.epa.gov/ednnrmrl/repository/cross/cross.pdf. An
update of this manual has been funded by EPA and will be published in the near future. It is a
collaborative effort between Pitt and the Center for Watershed Protection. The new manual will
include information on optical brightener monitoring, a quick and effective way for screening
large watersheds for illicit wastewater connections.
Procedures for identifying potential illicit discharges to storm drain systems include
reviewing existing drainage area maps, surveying building storm drain connections, and
inspecting sewer lines (U.S. EPA, 2003b). Visible flow during dry periods is a sign of a possible
cross connection that should be further investigated. Visual inspection of the insides of a sewer
system can be done with television equipment. Differences between known connections shown
on maps and those revealed by the television should be further investigated. Tracers are often
used to investigate suspected illicit connections (Pitt et al., 1993). A tracer is a parameter not
characteristic of the base flow; the particular tracer present is dependent on the content of the
illicit discharge. Tracers include water temperature, specific conductivity, fluoride and/or
hardness, ammonia and/or potassium, surfactants and/or fluorescence (including optical
brighteners from laundry detergents), chlorine, color, odor, turbidity, and flotables. Tracers for
microbial contamination would include sanitary wastewater parameters such as BOD or
suspended solids. Tracers can also be artificial, such as a dye. Smoke testing is another
investigative method for illicit connections. Zinc chloride smoke injected into the sewer lines
emerges from all breaks in the sewer line, vents in connected buildings, and outfalls (U.S. EPA,
2003b).
3.3.3 Nonstructural BMPs
Nonstructural BMPs include institutional and educational practices with the goal of
changing behaviors so that the amount of pollutants entering the stormwater drains and receiving
waters are reduced (Urbonas, 1999). These common sense measures for addressing microbial
contamination include limiting public and animal access to sensitive watershed or riparian areas,
public education on the role of storm drains, erosion control, vegetative buffers, street sweeping,
animal waste management, and pet waste or pooper-scooper ordinances. While quantitative data
on nonstructural BMP effectiveness are limited, a number of these practices have been shown to
reduce receiving-water bacteria levels in rural and agricultural settings, primarily by controlling
sources. They are provided here because some of the practices may apply to urban watersheds,
particularly developing rural areas. Several demonstrations are described in the report prepared
for EPA entitled Section 319 Nonpoint Source National Monitoring Program - Successes and
Recommendations (Lombardo et a/., 2000). The types of practices reported to be successful are
riparian/livestock exclusion fencing, riparian zone vegetation establishment and/or restoration,
improved grazing management including stream crossings, improved handling of barnyard
runoff and manure, campground educational programs on waste disposal, and upgrading septic
systems. Project updates included in the 2002 update report (Lombardo et al, 2002) available at
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http://h2osparc.wq.ncsu.edu/02rept319/indexframe.html show mixed results associated with
using BMPs for reducing nonpoint source microbiological contamination. Some of the relevant
results are presented below.
• The following BMPs were implemented in Arizona's Oak Creek Canyon Watershed:
erecting permanent barricades along a highway to significantly reduce visitor access to
the watershed's state park and campground, improving restroom facilities at the park and
campground, and educational outreach. While limited improvement to the water quality
in Oak Creek is attributed to these BMPs, the watershed task force is investigating
additional sources of fecal coliform that, if addressed, can result in further improvement.
• Reductions in fecal coliform in California's Morro Bay Watershed are attributed to
measures used to restrict or eliminate cattle access to riparian pastures.
• BMPs implemented in Washington's Totten and Eld Inlets are repair of failing on-site
wastewater treatment systems and implementation of farm plans on farms that potentially
threaten receiving-water quality. "Freshwater fecal coliform count and loading results
suggest that for Burns, Pierre, and McLane creeks, the degree of BMP installation and
maintenance is inadequate, and/or that unfactored demographic change may be eroding
what might otherwise be improved conditions. For Schneider and Perry creeks, where
water quality improved, the ability to link the improvement to pollution-control programs
is hampered by lack of a control in one case, by non-BMP land-use change in the other
case, and by inadequate BMP data in both cases. If effectiveness is measured by
significant lasting decreases in pollution, then the results allow the possibility of
effectiveness in these two cases. In those cases where pollution decreased, it appears to
be on the rise again, which suggests that nonpoint pollution-control programs need to be
at least cyclical if not continuous." (Batts and Seiders, 2003).
• A system of BMPs designed to exclude livestock from critical areas of streams and
riparian zones has contributed to a reduction in indicator bacteria counts from 29 to 40%
in Vermont's Lake Champlain Basin Watershed. Indicator bacteria counts exhibited
pronounced seasonal cycles - low in winter and high in the growing season beginning in
May. Additional experiments confirmed that indicator bacteria survive in stream
sediments during the warmer months and can be resuspended when the sediments are
disturbed. Decreases in E. coli and fecal coliform occurred during all seasons in the two
watersheds studied, while fecal streptococcus decreases were significant in one of the
watersheds. (Meals et a/., 2001).
• Erosion control and animal waste management practices were implemented in Alabama's
Lightwood Knot Creek Watershed. Although water quality improved for a number of
characteristics, fecal bacterial concentrations were not improved. Fecal coliform
concentrations decreased to some extent, but not to a significant degree. Fecal
streptococcus concentrations increased in the watershed. The relatively small change
was attributed to a design flaw in the constructed cattle crossing that encourages cows to
congregate on the crossing during dry periods. (Cook and O'Neil, 2003).
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• BMPs were shown to decrease indicator bacteria concentrations in North Carolina's
Long Creek Watershed. The 70% decrease in median fecal coliform levels in one part of
the creek is attributed to livestock exclusion. The installation of exclusion fencing in the
pasture of the area's largest dairy farm is believed to be responsible for 90% and 80%
decreases in fecal coliform and fecal streptococci levels.
Aside from farm animals, indigenous wildlife, rodents, and pets can increase indicator
microorganism concentrations to levels that exceed water quality standards. In Northern
Virginia's Four Mile Run Watershed, microbial source tracking identified a number of species
(waterfowl, raccoon, human, dog, deer, and Norway rat) as the E. coli sources (Simmons, Jr. et
a/., 2000; NVRC, 2002). The TMDL developed for fecal coliform requires that loadings from
waterfowl, raccoon, dog and other wildlife, as well as humans, be reduced by significant
percentages (NVRC, 2002). Although nonstructural BMPs will likely be used, the TMDL
document does not address how achieving the TMDL goal will be accomplished. The approach
will be presented in the TMDL implementation plan to be developed.
Instituting pet waste management or pooper-scooper laws is the traditional way
communities have dealt with pet waste, which can contaminate water bodies or pose a potential
threat to residents through direct contact. Waye (2003) cites the success of dog parks as BMPs.
These parks should be located away from water bodies, and provide fencing, public education on
managing waste, and disposal bags and receptacles. Having a local community pet group take
responsibility for a park and establishing the norm of picking up after one's own pets help to
ensure success of these parks.
Other nonstructural BMPs include modifying storm drain inlets to impede rodent access,
public education, labeling storm drain inlets, and street sweeping.
3.3.3.1 Managing Waste from Resident Canada Geese
In recent years, geese populations have grown in many areas in the U.S. The problems
encountered by local communities are the health and cosmetic problems associated with the fecal
material generated, as well as the number of geese, and related traffic and safety concerns as
these large birds cross traffic. Municipalities are instituting measures to protect public health
from the impacts associated with this waste. The coastal town of Spring Lake, in New Jersey's
Monmouth County, is experiencing high bacteria levels in a pond occupied by many Canada
geese. During rain events, the pond overflows into the ocean, resulting in beach closures. The
municipality automatically bans swimming at the nearby ocean beaches for 24 hours after it
rains at least one-tenth of an inch (Bates, 2003). Restricting contact with recreational waters
during wet-weather events is practiced by many municipalities as a precautionary measure
because of the potential for waterborne illnesses to result in swimmers in contact with pathogens
in the wet weather discharges.
Colts Neck, another Monmouth County community, recently contracted with the U.S.
Department of the Interior's Fish and Wildlife Service to asphyxiate Canada geese at local
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ponds. The local health officer defended the action based on the nuisance and potential health
hazards posed by the geese droppings in and around the ponds (Jordan, 2003). Some citizens
and animal rights advocates opposed the action and proposed alternatives. Waye (2003) names
the possible alternatives identified by GeesePeace (www.geesepeace.org), including egg addling,
vegetative barriers around water bodies, border collie patrols, goose repellants, and "no feed"
zones.
Most Canada geese populations are migratory, wintering in the U.S. and migrating north
to summer breeding grounds in the Canadian Arctic. The availability of park-like open spaces
with short grass adjacent to water bodies have resulted in growing numbers of locally-breeding
geese in the U.S. known as resident Canadian geese. There are an estimated 3.5 million resident
Canada geese in the U.S. (U.S. Fish and Wildlife Service, 2002). Resident geese are protected
under the Migratory Bird Treaty Act of 1918 and the Migratory Bird Conservation Act of 1929,
and cannot be legally taken during a hunting season, unless a special federal permit is obtained
from the Service. The proposed draft Environmental Impact Statement (EIS) released March 4,
2002, by the U.S. Fish and Wildlife Service grants the States the authority to implement
approved population control strategies, such as nest and egg destruction, and trapping and
culling programs, without having to go through the permit process. Until the draft EIS is
finalized, scheduled for the fall 2003, states must obtain a special permit from the Service for
resident Canada geese population control strategies.
3.3.4 Effects ofBMPs on Receiving-Water Quality
From the available information on structural and nonstructural BMPs, it is evident that
more research is needed on their effectiveness in reducing microbiological loads in stormwater
runoff. Further, there should be a distinction between the effectiveness of structural and
nonstructural BMPs. The highly variable effectiveness data exhibited by structural BMPs
indicate that a variety of conditions affect the behavior of microorganisms and thus performance.
These include BMP volume, temperature, light intensity, wetland plant type, filter design, and
maintenance scope and frequency. As research in these areas progresses, BMP designs and
O&M requirements can be aimed at achieving improved results. With even less quantitative
information available for nonstructural BMPs, studies of their effectiveness in watersheds will
provide information for health and environmental managers in other watersheds.
A concern with using BMPs to treat stormwater is that the microbial densities in the
effluents may exceed water quality standards, even in BMPs considered to be performing well.
For example, Davies and Bavor (2000) report a geometric mean for Enterococcus concentrations
of 9.0 x 102/100 mL for the wetland's outflow, which is much higher than the U.S. recreational
fresh water standard of 33/100 mL. In a case like this, the receiving water will need to have a
high enough flow rate or volume to achieve the water quality target through dilution. Therefore,
using a single BMP may not provide the level of treatment needed, in which case other options
will need to be considered. These include incorporating a preliminary treatment step upstream
of a structural BMP to create a treatment train or disinfecting the stormwater. Reducing runoff
volume and source control are the most reliable ways to decrease microorganism loads to
receiving waters from stormwater.
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3.4 Conclusions
Managing microbial contamination in urban watersheds presents unique challenges. A
primary reason for this is that some of the microorganism content in runoff and waterways
occurs naturally because microorganisms are components of waste products deposited by
animals residing in these watersheds. The populations of these organisms vary with animal
population and are affected by environmental factors such as temperature, sunlight, and nutrient
availability. Also, quantitative effectiveness results of the BMPs used to manage this diffuse
source pollution are often unavailable or inconclusive. Therefore, managers relying on BMPs
for allocating nonpoint source loads to achieve a TMDL goal need to be prepared to revise
management plans and even allocations if monitoring data reveals that the desired results are not
achieved.
Although some quantitative information on the effectiveness of structural BMPs for
managing microbial contamination in stormwater is available, the amount of information is less
plentiful than it is for other contaminants. Microorganism or pathogen removal has been
reported most frequently for sand filters, wetlands, and wet detention ponds. However, the
results are highly variable. The available wet pond fecal coliform data shows removals between
46 and 99 percent, except for one site where the removal was -5.8%. The wetlands efficiency
data reviewed has an even greater range of removal efficiencies, from -134% to 97%. These
results contradict some research studies with findings that show wetlands have better removal
efficiencies than ponds. Research to understand the key biological, chemical, and physical
processes controlling microorganism behavior in commonly used stormwater BMPs is necessary
(Sullivan and Borst, 2001). This research would better define the relationships between design
parameters and effectiveness and will contribute to the development of models that will predict
effluent quality over a BMP's lifetime, temporal variations of effluent quality, and differences in
performance due to differences in events (Sullivan and Borst, 2001). Also useful would be
increased understanding of the relationships between common water quality parameters, e.g.,
TSS, and microbial indicators and pathogens.
Less quantitative information is available on the effectiveness of nonstructural BMPs
than on the structural BMPs discussed above. EPA's Nonpoint Source National Monitoring
Program generates some data that shows decreases, increases, and no change after BMP
implementation. The watersheds described in the available program summaries are primarily
rural in nature. Public education and pet waste management regulations and programs are other
nonstructural BMPs that show promise for urban watershed management but for which
quantitative performance data are needed.
Disinfection of CSO and other WWF types achieves a much greater degree of
microorganism removal than BMPs. It's also been the subject of a much greater amount of
research and investigation. Disinfection has been demonstrated to reduce microorganism
concentrations in WWFs with high concentrations (105 to 107 organisms/100 mL) by several
orders of magnitude and produce effluents meeting permit discharge requirements (102 to 103
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organisms/100 mL). WWF disinfection generally occurs within shorter contact times than
conventional wastewater disinfection, i.e., less than 30 minutes, with intense mixing to ensure
disinfectant contact with the maximum number of microorganisms, and increased disinfectant
dosage. Effective use of this high-rate disinfection process requires use of a treatment train, with
an initial treatment of either filtration or inertial separation (e.g., sedimentation and vortex) to
remove suspended solids. This is to address the phenomena of microbial aggregation and
particle association/occlusion that cause decreased disinfection efficiency.
Chlorination is the only chemical disinfection technology currently used for disinfection
of WWF. Although effective, this technology generates formation of chlorinated hydrocarbons,
i.e., chloroform and THMs, which are suspected carcinogens. To address this concern and
remove excess free C12, the chlorination process can be augmented by dechlorination with either
gaseous sulfur dioxide or sodium bisulfite solution. Other disinfection technologies investigated
for CSO include UV light irradiation, C1O2, and O3. Of these three technologies, only UV
disinfection has recently entered commercial use for WWF disinfection. Chlorine dioxide and
O3 have not been put to commercial use in the U.S. Removal efficiencies for the disinfection
technologies discussed (C12, UV, C1O2, and O3) achieve bacterial reductions of 99.9% to
99.99%. This is a significantly greater level of contaminant reduction than is achieved by
BMPs. Although just beginning to be used for treating stormwater, disinfection of stormwater
may be necessary to achieve water quality objectives in some watersheds.
A final point that should be considered is the uncertainty associated with the use of
indicator microorganisms to determine pathogen reductions resulting from the use of a control
technology or a BMP. Chapter 1 explores the relationships between indicators, pathogens, and
waterborne illness. Although the desired reduction of an indicator microorganism density,
TMDL, or water quality target is achieved by a certain technology or a management approach,
there is still a possibility of public health impact due to the presence of disease-causing
microorganisms, i.e., pathogens. Alternatively, the indicators may have provided a false or
exaggerated indication of the presence of disease-causing pathogens and, thus, no benefit to
human health was achieved through the control or management practice implemented.
Watershed managers need to be aware of the limitations associated with indicators and
remember the primary goal of protecting public health.
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