Guidelines for Canadian Recreational Water Quality

Guidelines for Canadian Recreational Water Quality
Guidelines for
Canadian Recreational
Water Quality
Third Edition
Health Canada is the federal department responsible for helping the people of Canada
maintain and improve their health. We assess the safety of drugs and many consumer
products, help improve the safety of food, and provide information to Canadians to help
them make healthy decisions. We provide health services to First Nations people and to
Inuit communities. We work with the provinces to ensure our health care system serves the
needs of Canadians.
Published by authority of the
Minister of Health.
Guidelines for Canadian Recreational Water Quality
is available on Internet at the following address: http://www.healthcanada.gc.ca/
Également disponible en français sous le titre :
Recommandations au sujet de la qualité des eaux utilisées à des fins récréatives au
Canada
This publication can be made available on request
in a variety of alternative formats.
For further information or to obtain additional copies, please contact:
Publications
Health Canada
Ottawa , Ontario K1A 0K9
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E-Mail: [email protected]
© Her Majesty the Queen in Right of Canada, represented by the Minister of Health, 2012
This publication may be reproduced without permission provided the source is fully acknowledged.
Cat.: H129-15/2012E
ISBN: 978-1-100-20892-3
Guidelines for
Canadian Recreational Water
Quality
Third Edition
Prepared by the
Federal-Provincial-Territorial Working Group
on Recreational Water Quality
of the
Federal-Provincial-Territorial Committee on
Health and the Environment
Ottawa, Ontario
April, 2012
This document may be cited as follows:
Health Canada (2012). Guidelines for Canadian Recreational Water Quality, Third Edition.
Water, Air and Climate Change Bureau, Healthy Environments and Consumer Safety Branch,
Health Canada, Ottawa, Ontario. (Catalogue No H129-15/2012E).
The document was prepared by the Federal-Provincial-Territorial Working Group on
Recreational Water Quality of the Federal-Provincial-Territorial Committee on Health and the
Environment.
Any questions or comments on this document may be directed to:
Water, Air and Climate Change Bureau
Healthy Environments and Consumer Safety Branch
Health Canada
269 Laurier Avenue West, Address Locator 4903D
Ottawa, Ontario
Canada K1A 0K9
Tel.: 613-948-2566
Fax: 613-952-2574
E-mail: [email protected]
Website: www.healthcanada.gc.ca/waterquality
Table of Contents
Members of the Federal-Provincial-Territorial Working Group on Recreational Water
Quality............................................................................................................................................ 1
Acknowledgements ....................................................................................................................... 2
Executive Summary ...................................................................................................................... 3
Introduction ................................................................................................................................... 8
Part I: Management of Recreational Waters ........................................................................... 10
1.0
The multi-barrier approach ........................................................................................... 11
2.0
Environmental Health and Safety Survey .................................................................... 12
2.1
Pre-survey preparations ........................................................................................ 12
2.2
On-site visit ........................................................................................................... 13
2.3
Assessment report ................................................................................................. 13
3.0
Barriers for protection of recreational waters ............................................................. 15
3.1
Compliance monitoring ........................................................................................ 15
3.1.1 Frequency of microbiological sampling ................................................... 16
3.1.2 Location of microbiological sampling ...................................................... 18
3.1.3 Other monitoring tools .............................................................................. 19
3.2
Public awareness and communication .................................................................. 21
3.3
Public health advice .............................................................................................. 23
3.4
Hazard control actions .......................................................................................... 24
Part II: Guideline Technical Documentation ........................................................................... 25
4.0
Recommended indicators of faecal contamination ...................................................... 25
4.1
Indicator organisms for primary contact recreation .............................................. 26
4.1.1 Fresh waters: Escherichia coli (E. coli) .................................................... 26
4.1.2 Marine waters: Enterococci ...................................................................... 34
4.2
Advice regarding water intended for secondary contact recreational activities ... 40
4.3
Other organisms as potential indicators ................................................................ 42
5.0
Pathogenic microorganisms ........................................................................................... 52
5.1
Bacterial pathogens ............................................................................................... 52
5.1.1 Enteric pathogenic bacteria (Campylobacter, pathogenic E. coli,
Salmonella, Shigella) ................................................................................ 53
5.1.2 Free-living pathogenic bacteria (Aeromonas, Legionella, Pseudomonas,
Mycobacterium) ........................................................................................ 57
5.1.3 Other pathogenic bacteria (Leptospira, Staphylococcus) ......................... 61
5.2
Viral pathogens ..................................................................................................... 63
5.2.1 Enteric viruses ........................................................................................... 64
5.3
Protozoan pathogens ............................................................................................. 68
5.3.1 Enteric protozoa (Giardia, Cryptosporidium) ........................................... 68
5.3.2 Free-living protozoa .................................................................................. 71
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Guidelines for Canadian Recreational Water Quality: Third Edition
6.0
Cyanobacteria and their toxins...................................................................................... 73
7.0
Other biological hazards ................................................................................................ 88
7.1
Schistosomes causing swimmer’s itch .................................................................. 88
7.2
Aquatic vascular plants and algae ......................................................................... 92
7.3
Additional organisms ............................................................................................ 93
8.0
Physical, aesthetic and chemical characteristics .......................................................... 93
8.1
Physical characteristics ......................................................................................... 94
8.1.1 pH.............................................................................................................. 94
8.1.2 Temperature .............................................................................................. 95
8.2
Aesthetic characteristics........................................................................................ 97
8.2.1 Turbidity ................................................................................................... 97
8.2.2 Clarity (light penetration) ......................................................................... 99
8.2.3 Colour ....................................................................................................... 99
8.2.4 Oil and grease ......................................................................................... 101
8.2.5 Litter ........................................................................................................ 102
8.3
Chemical characteristics ..................................................................................... 102
9.0
Faecal contamination and beach sand ........................................................................ 104
10.0
Faecal pollution source tracking.................................................................................. 109
References .................................................................................................................................. 112
Appendix A: List of acronyms and abbreviations ................................................................ 135
Appendix B:
B.1
B.2
B.3
B.4
Microbiological sampling and analysis ........................................................... 136
Sampling procedures for water ........................................................................... 136
Sampling procedures for sand and sediments ..................................................... 137
Sample transport, preservation and storage ........................................................ 137
Methods for microbiological analysis ................................................................. 138
B.4.1 Recommended indicators of faecal contamination ................................. 138
B.4.2 Pathogenic microorganisms .................................................................... 140
Appendix C:
C.1
C.2
C.3
C.4
Composite sampling for faecal contamination ............................................... 145
Description .......................................................................................................... 145
Introduction ......................................................................................................... 145
Study results ........................................................................................................ 146
Conclusions ......................................................................................................... 147
Appendix D: Recreational swimming area Environmental Health and Safety Survey
(EHSS) checklist ........................................................................................................................ 148
Appendix E: Example: Sample collection and reporting form ........................................... 153
Appendix F: Examples: Informative beach signs ................................................................. 154
F.1
Beach posting: example of an informative swimming advisory sign ................. 154
F.2
Beach posting: example of an informative water suitable for swimming sign ... 155
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Guidelines for Canadian Recreational Water Quality: Third Edition
Members of the Federal-Provincial-Territorial Working Group on
Recreational Water Quality
Provincial Members
Alberta
Karina Bodo
Public Health Surveillance and
Environmental Health
Alberta Health and Wellness
British Columbia
Howard Singleton
Population Health and Wellness
British Columbia Ministry of Health
Manitoba
Dwight Williamson (Chair)
Water Science and Management Branch
Manitoba Water Stewardship
New Brunswick
Diane Fury
Public Health and Medical Services Division
New Brunswick Department of Health and
Wellness
Newfoundland and Labrador
Renee Paterson
Water Resources Management Division
Newfoundland and Labrador Department of
Environment and Conservation
Newfoundland and Labrador (Alternate)
Ali Khan
Water Resources Management Division
Newfoundland and Labrador Department of
Environment and Conservation
Nova Scotia
Darrell Taylor
Water and Wastewater Branch
Nova Scotia Department of the Environment
Quebec
Denis Brouillette
Direction du suivi de l’état de
l’Environnement
Ministère du Développement durable, de
l’Environnement et des Parcs du Québec
Ontario
Tim Fletcher
Standards Development Branch
Ontario Ministry of the Environment
Ontario (Alternate)
Cassandra Lofranco
Standards Development Branch
Ontario Ministry of the Environment
Corresponding Members
Prince Edward Island
Joe Bradley
Environmental Health Services
Prince Edward Island Department of Health
Saskatchewan
Tim Macaulay
Population Health Branch
Saskatchewan Health
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Guidelines for Canadian Recreational Water Quality: Third Edition
Invited Members
Environment Canada
Tom Edge
National Water Research Institute
Environment Canada
U.S. Environmental Protection Agency
Alfred Dufour
National Exposure Research Laboratory
United States Environmental Protection
Agency
Secretariat
Health Canada
Will Robertson
Healthy Environments and Consumer Safety
Branch
Health Canada
Health Canada
Gordon Yasvinski
Healthy Environments and Consumer Safety
Branch
Health Canada
Acknowledgements
The contributions of the following individuals to the development of these guidelines are also
acknowledged:
Justin Balko
Alexandra Bourne
Cecily Flemming
Julie Kinzelman
Debra Mooney
Wendy Ralley
Alberta Health and Wellness
Manitoba Water Stewardship
Ontario Ministry of the Environment
City of Racine (Wisconsin) Health Department
Alberta Health and Wellness
Manitoba Water Stewardship
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Guidelines for Canadian Recreational Water Quality: Third Edition
Executive Summary
The primary goal of the Guidelines for Canadian Recreational Water Quality is the protection of
public health and safety. This document provides guidance on the factors that can interfere with
the safety of recreational waters from a human health perspective. It is intended to guide
decisions by provincial and local authorities that are responsible for the management of
recreational waters.
Recreational water quality generally falls under provincial and territorial jurisdiction.
Responsibility for the safe management of recreational waters can be shared between the
provincial-territorial authorities and the beach managers or service providers. The division of
duties (e.g. responsibility for monitoring or the communication of results) may vary depending
on provincial-territorial policies in place. Recreational waters are considered to be any natural
fresh, marine or estuarine bodies of water that are used for recreation. These include lakes and
rivers, as well as human-made constructions (e.g. quarries, artificial lakes) that are filled with
untreated natural waters. The principal health risk associated with exposure to recreational water
quality hazards is infection as a result of contact with pathogenic microorganisms. Other risks
include injury or illness due to the physical or chemical properties of the water.
The Guidelines for Canadian Recreational Water Quality consider the human health risks
associated with recreational activities—primary contact activities, such as swimming (this
includes bathing/wading for the purposes of this document), windsurfing and waterskiing, as
well as secondary contact activities, such as canoeing or fishing—in natural waters through
intentional or incidental immersion. They establish guideline values for specific parameters used
to monitor recreational water quality, including the bacteriological indicators of faecal
contamination, cyanobacteria and their toxins, and values for physical and aesthetic objectives.
This document also outlines a risk management approach to safe recreational water quality and
describes the current scientific knowledge regarding the water quality hazards that can be
encountered in the natural recreational water environment. It discusses pathogenic
microorganisms of concern, water sampling and analysis, as well as emerging issues, such as
faecal contamination of beach sand and faecal pollution source tracking.
Management of recreational waters
The protection and safe management of recreational waters require the cooperation of all
stakeholders. The best approach is based on a preventive risk management strategy that focuses
on the identification and control of water quality hazards and their associated risks before users
could be exposed. As with drinking water, the multi-barrier approach provides this preventive
strategy through an integrated system of procedures, actions and tools that collectively reduce the
risk of human exposure to recreational water quality hazards. The effectiveness of these
procedures, actions and tools is then verified or confirmed by monitoring results and the
application of guideline values. The success of this approach rests primarily with the
establishment of multiple barriers to protect watersheds.
Potential water quality hazards or risk scenarios that can affect the recreational water area need
to be identified through an Environmental Health and Safety Survey. The results of this survey
are then used to identify the appropriate procedures or actions that should be put in place as
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Guidelines for Canadian Recreational Water Quality: Third Edition
barriers. These may include physical actions, such as beach cleanup and grooming, or processes
or tools to improve the effectiveness of the recreational water management program, such as
monitoring, guidelines and standards, and education and communication strategies.
Guideline values and technical information
Guideline values for a variety of water quality parameters are one important component of the
overall risk management approach to safe recreational water quality. They should be used
together with the appropriate technical documentation provided for these parameters. Table 1
outlines the guideline values for the recommended water quality parameters. Considerations are
also provided for parameters and water quality hazards for which guideline values cannot be
established.
Table 1. Guidelines for Canadian recreational water quality: summary table
Guidelines
Parameter
Escherichia coli
(Primary-Contact
Recreation) ∗
Enterococci
(Primary-Contact
Recreation) ∗
Pathogenic Microorganisms
(bacteria, viruses, protozoa)
Cyanobacteria
Cyanobacterial toxins
Other Biological Hazards
(e.g. schistosomes causing
swimmer’s itch; aquatic
vascular plants and algae)
pH
Temperature
Chemical Hazards
Considerations
Guideline
Geometric mean concentration (minimum 5
samples)
Single sample maximum concentration
≤ 200 E. coli /100 mL
≤ 400 E. coli /100 mL
Geometric mean concentration (minimum 5
samples)
Single sample maximum concentration
≤ 35 Enterococci /100 mL
≤ 70 Enterococci /100 mL
Testing only needed when there is epidemiological
or other evidence to suggest that this is necessary
Total Cyanobacteria
Total Microcystins
Recreational activities should not be pursued in
waters where the responsible authority deems the
presence of these organisms poses a risk to the health
and safety of the users
For waters used for primary contact recreation
Should not cause an appreciable increase or decrease
in the deep body temperature of swimmers
Risks associated with specific chemical hazards will
be dependent on the particular circumstances of the
area and should be assessed on a case-by-case basis.
No numerical guideline value
≤ 100,000 cells/mL
≤ 20 μg/L
No numerical guideline value
5.0 to 9.0
No numerical guideline value
No numerical guideline value
Aesthetic Objectives
Parameter
Turbidity
Clarity
Colour
∗
Considerations
Aesthetic Objective
To satisfy most recreational uses
Clarity should be sufficient for users to estimate
depth and to see subsurface hazards
Colour should not be so intense as to impede
visibility in areas used for swimming
50 NTU
Secchi Disc visible at a depth
of 1.2 m
No numerical value
Advice regarding waters intended for secondary-contact recreational activities is provided in Section 4.2.
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Guidelines for Canadian Recreational Water Quality: Third Edition
Oil and Grease
Litter
Should not be present in concentrations that can be
detected as a visible film, sheen, discolouration or
odour; or that can form deposits on shorelines or
bottom sediments that are detectable by sight or
odour
Areas should be free from floating debris as well as
materials that will settle to form objectionable
deposits
No numerical value
No numerical value
Indicators of faecal contamination—primary contact recreation
Escherichia coli is the most appropriate indicator of faecal contamination in fresh recreational
waters, and enterococci is the most appropriate indicator of faecal contamination in marine
recreational waters. Guideline values for E. coli and enterococci have been developed based on
the analysis of epidemiological evidence relating concentrations of these organisms to the
incidence of swimming-associated gastrointestinal illness observed among swimmers. The
values represent risk management decisions based on the assessment of possible health risks for
the recreational water user and the recognition of the significant benefits that recreational water
activities provide in terms of health and enjoyment. The guidelines advised that recreational
water areas routinely used for primary contact recreation be monitored at a minimum of once per
week, with increased monitoring recommended for those beaches that are highly frequented or
are known to experience high user densities. Similarly, under certain scenarios, a reduction in the
recommended sampling frequency may be justified.
Advice regarding water intended for secondary-contact recreational activities
Due to increased interest from jurisdictions in distinguishing between primary contact activities
and secondary contact activities, this current edition of the Guidelines takes an initial step at
providing advice for secondary contact activities and faecal indicator concentration. There are
insufficient epidemiological data available to derive precise health-based faecal indicator limit
values intended to protect users engaged in secondary contact recreational activities from
exposure to faecal contamination. Secondary contact is defined as recreational activity in which
only the limbs are regularly wetted and in which greater contact (including swallowing water) is
unusual. Because a lower degree of water exposure can be expected at most times during the
majority of secondary contact recreational activities, there may be some waters in which a
secondary contact use designation with separate water quality values is desired and considered
acceptable to management and regulatory authorities. Advice is provided that the application of a
factor of 5 to the existing geometric mean faecal indicator concentration used to protect primary
contact recreation users may be used as an approach to establish faecal indicator limits. These
values represent a risk management decision based on the assessment of the expected exposure
scenarios and potential health risks for the recreational water user, and represents a tolerable and
reasonable approach to protecting users engaged in a voluntary activity.
Other potential indicator organisms
The organisms most widely discussed as potential recreational water indicators include
Bacteroides spp., Clostridium perfringens, F+ RNA coliphages and bacteriophages infecting
Bacteroides fragilis. At present, none of these organisms meets a sufficient number of the
requirements necessary to be successfully used as a routine indicator of recreational water
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Guidelines for Canadian Recreational Water Quality: Third Edition
quality. These organisms appear to be better suited as possible pathogen indicators or as faecal
source indicators. Advances in detection and enumeration methods may improve the
understanding of these organisms and the roles they may play in future recreational water
monitoring programs.
Pathogenic microorganisms (bacteria, viruses, protozoa)
The challenges associated with the detection of pathogenic microorganisms in recreational
waters are currently too great to recommend this practice as part of a regular monitoring
program. Surveillance is necessary only during special circumstances, such as during waterborne
disease outbreak investigations. Faecal indicators such as E. coli and enterococci are the best
available indicators for the possible presence of enteric pathogenic microorganisms. However,
the absence of the recommended faecal indicators should not be interpreted to mean that all
pathogenic microorganisms are also absent.
Cyanobacteria and their toxins
Swimmer illnesses have been reported following exposure to toxic cyanobacterial blooms in
recreational waters. Guideline values for cyanobacteria and their toxins (microcystins) have been
established to protect against both the risk of exposure to microcystins as well as any harmful
effects that may be possible as a result of exposure to high densities of cyanobacterial material.
Waters shown to exceed the established guideline values or those in which a bloom has
developed may result in human exposure to cyanobacterial material or cyanotoxins in amounts
sufficient to be harmful to human health. A swimming advisory may be issued at the discretion
of the responsible authority. Contact with waters where an advisory has been issued should be
avoided until the advisory has been rescinded.
Other biological hazards
Recreational water activities should not be pursued in areas where other biological hazards are
present in significant quantities such that they pose a risk to the health and safety of recreational
water users. Examples include the presence of organisms responsible for swimmer’s itch and
dense growths of aquatic plants.
Physical, aesthetic and chemical characteristics
Physical, aesthetic and chemical characteristics of water can have an impact on recreational
water users. Recreational waters should be of good aesthetic quality and should be free from
substances that impair its aesthetic appreciation. Aesthetic components can also impact the
health and safety of recreational water users where visibility has become significantly impaired.
Guideline values for specific chemical parameters in recreational waters cannot be specified.
In general, potential risks from exposure to chemical parameters will site specific and be much
smaller than microbiological risks. It is important for beach operators or service providers to
have a mechanism in place to ensure that risks from potential chemical hazards are identified and
adequate action is taken.
Faecal contamination of beach sand
Beach sand can present an important non-point source of faecal contamination to recreational
waters. Sand may provide a favourable environment for microorganisms of faecal origin,
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Guidelines for Canadian Recreational Water Quality: Third Edition
permitting them to survive for longer periods than in the adjacent waters. Physical factors such
as wave action, storm surges, tidal activity and high swimmer load can result in the transference
of faecal microorganisms from foreshore and nearshore sand and sediments to waters used for
swimming.
Further research is needed to determine the relationships between faecal indicator bacteria and
the possible presence of faecal pathogens in beach sand, as well as the potential implications for
human health. Barriers that collectively reduce the risk of exposure for beach users could include
public education campaigns, improved beach sanitation practices, appropriate sand grooming
practices and actions designed to discourage the activities of animals (birds and other wildlife)
within the beach area.
Faecal pollution source tracking
Faecal pollution source tracking is an emerging field that focuses on understanding the specific
sources of faecal contamination affecting an area. Numerous chemical and microbiological
source-tracking tools have been described. A good understanding and formulation of the nature
of the faecal contamination problem are required before any faecal source tracking study can be
considered.
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Guidelines for Canadian Recreational Water Quality: Third Edition
Introduction
Recreational water quality generally falls under provincial and territorial jurisdiction.
Responsibility for the safe management of recreational waters can be shared between the
provincial or territorial authorities and the beach managers or service providers. The division of
duties (e.g. responsibility for monitoring or the communication of results) will vary depending
on provincial or territorial policies in place. The Federal-Provincial-Territorial Working Group
on Recreational Water Quality was established by the Federal-Provincial-Territorial Committee
on Health and the Environment to review and evaluate current scientific information on
recreational water quality and develop up-to-date guidance. This has resulted in the development
of an updated version or Third Edition of the Guidelines for Canadian Recreational Water
Quality, which incorporates current science and outlines a recommended risk management
approach. In preparing this document, the Working Group re-evaluated the criteria for existing
indicators of recreational water quality and conducted reviews of the literature published on the
topic of recreational water quality and human health and safety, including research papers,
reports of epidemiological investigations, published texts, disease surveillance reports and
guideline documentation developed by other government and multinational organizations
worldwide.
The primary goal of the Guidelines for Canadian Recreational Water Quality is the protection
of public health and safety. The document is aimed primarily at responsible authorities and
decision-makers. It provides guidance on the factors that can interfere with the safety of
recreational waters from a human health perspective. It recommends the adoption of a preventive
risk management strategy that focuses on the identification and control of water quality hazards
prior to the point of contact with the recreational water user. It also recommends the use of a
multi-barrier approach as the most effective means for protecting users from exposure to water
quality hazards in recreational waters.
Recreational waters can be considered as any natural fresh, marine or estuarine bodies of water
where a significant number of people use the water for recreation. These include human-made
constructions using untreated natural waters (e.g. artificial lakes, quarries).
Recreational water activity can be classified as any activity involving intentional or incidental
immersion in natural waters. These are further defined (adapted from WHO, 2003a) as follows:
•
Primary contact: Activities in which the whole body or the face and trunk are frequently
immersed or the face is frequently wetted by spray, and where it is likely that some water
will be swallowed (e.g., swimming, surfing, waterskiing, whitewater
canoeing/rafting/kayaking, windsurfing, subsurface diving).
.
•
Secondary contact: Activities in which only the limbs are regularly wetted and in which
greater contact (including swallowing water) is unusual (e.g., rowing, sailing, canoe touring,
fishing).
This document does not include treated recreational water facilities (e.g., swimming pools, hot
tubs, whirlpool baths, hydrotherapy pools) or tertiary contact uses of water, where no contact
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Guidelines for Canadian Recreational Water Quality: Third Edition
with water is expected (e.g., walking along the shore, sunbathing). It does not address waterrelated injuries such as drowning or diving injuries. It does not address issues specific to
particularly sensitive individuals or population groups. Individuals concerned about their health
status or the health of vulnerable population groups should consult with their health care
provider or regional health unit in order to make an informed decision before engaging in any
recreational water activities.
The document is divided into two parts:
•
Part I (Management of Recreational Waters) provides guidance on the management of
recreational waters, including approaches for water quality hazard assessment, water quality
monitoring and the implementation of preventive or corrective actions.
•
Part II (Guideline Technical Documentation) establishes guideline values and aesthetic
objectives and provides related technical and scientific information on the water quality
parameters and hazards of importance for Canadian recreational waters.
The guideline values and aesthetic objectives established in this document should not be
regarded as legally enforceable standards, except where adopted by the appropriate
provincial/territorial or federal agency. Further, the jurisdictional authority may wish to apply
more stringent values and objectives as deemed necessary. This document is intended to guide
authorities responsible for developing operational standards as part of a comprehensive beach
management plan. The Guidelines for Canadian Recreational Water Quality may be periodically
revised or adjusted as necessary to ensure that they continue to remain protective of the health
and safety of all Canadians.
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Guidelines for Canadian Recreational Water Quality: Third Edition
Part I: Management of Recreational Waters
The division of duties pertaining to the safe management of recreational waters will differ
depending on the provincial or territorial policies in place. The authority overseeing the day-today operations of the recreational water area generally possesses the most comprehensive
knowledge of the area and is therefore in the best position to take the actions necessary to ensure
the safe operation of the facilities. The management information in this section is more pertinent
to managed beaches (either public or private); however the same principles can be applied to any
natural recreational water area.
Effective recreational water management requires the cooperation of all of its stakeholders,
including beach operators and service providers, governments, local businesses and industry, as
well as users. All stakeholders are expected to become informed about their roles and
responsibilities in the safe management of recreational waters.
A preventive multi-barrier approach to management that focuses on the identification and control
of water quality hazards and their associated risks before the point of contact with the
recreational water user represents the best strategy for the protection of public health from risks
associated with recreational waters. Reactive management strategies relying on compliance
monitoring alone will not be sufficient in protecting the health of the recreational water user.
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Guidelines for Canadian Recreational Water Quality: Third Edition
1.0
The multi-barrier approach
The multi-barrier approach is an integrated system of procedures, actions and tools that
collectively reduce the risk of human exposure to recreational water quality hazards. The concept
is analogous to the “source to tap” approach used for the management of safe drinking water
supplies in Canada (CCME, 2004).
This approach to recreational water management has been recognized by water quality
professionals worldwide. The concepts of preventive risk management and the use of multiple
barriers were at the heart of the recommendations for improved management of recreational
waters proposed by an international panel of experts following a meeting in Stockholm, Sweden,
in 1999 sponsored by the World Health Organization (WHO). A report on the outcome of this
meeting was later published in the document that has come to be known as the “Annapolis
Protocol” (WHO, 1999). Similarly, it is this approach that formed the basis for the management
framework outlined in the Guidelines for Safe Recreational Water Environments (WHO, 2003a).
The multi-barrier approach achieves success by having numerous barriers in place across all
identified areas of management (e.g. source protection, monitoring, hazard control,
communication, consultation), rather than focusing all efforts on a single barrier.
Specific benefits include:
• more effective public health protection;
• improved recreational water management (operational plans can be specifically tailored to
address an area’s individual needs and resources);
• improved public communication (leading to better public understanding of key concepts and
the public’s role in ensuring recreational water safety); and
• better management of emergencies (potential water quality hazards are understood and plans
are in place to address the problems effectively).
The following sections provide additional information on the different elements of the multibarrier strategy, including situation assessment (Environmental Health and Safety Surveys) and
application/implementation of barriers (Compliance Monitoring, Public Awareness and
Communication, Public Health Advice, Hazard Control Actions).
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Guidelines for Canadian Recreational Water Quality: Third Edition
2.0
Environmental Health and Safety Survey
An Environmental Health and Safety Survey (EHSS) provides the foundation or “blueprint”
for designing and implementing an effective risk management plan for recreational waters. It
is a comprehensive search for, and assessment of, existing and potential water quality hazards
(biological, chemical and physical) and their associated risks to the health and safety of the
public at designated beach areas. The EHSS also represents a general review of all aspects of a
beach’s operation. The data collected provide beach operators, service providers and responsible
authorities with the information necessary to make sound risk management decisions and to
develop and maintain an effective beach monitoring program. The EHSS fits under a multibarrier approach to recreational water management by identifying priority areas for which
interventions can be applied to reduce the level of risk for recreational water users.
An EHSS should be conducted on an annual basis, just beefore the start of the swimming season.
This survey should:
• catalogue the recreational water area’s basic characteristics;
• identify any potential sources of faecal contamination;
• identify any other potential physical, chemical or biological water quality hazards or
potential sources of such that may present a risk to recreational water users; and
• evaluate the effectiveness of the monitoring programs and risk management measures
currently in place.
The authority with the best knowledge of the day-to-day operation of the beach is the likely
candidate to lead this process. The EHSS process can also benefit greatly from intersectoral
collaboration. Persons or groups valuable to consult on the process can include:
• the appropriate provincial or territorial management or regulatory authority;
• beach managers;
• public and environmental health departments;
• community members; and/or
• individuals representing local business and industry.
The EHSS process consists of three basic steps: pre-survey preparations, the on-site visit and the
assessment report.
2.1
Pre-survey preparations
The pre-survey preparation step involves the collection and review of any and all information
available on the beach and adjacent area, including reports of any previous surveys. It can
provide valuable information on historical trends, problems and successes, which will help
ensure a more thorough and efficient on-site visit. Initial preparations may begin with a review of
basic beach information, such as the beach’s physical characteristics, the types of activities
practised and estimates of beach attendance. The use of topographical maps, aerial photos and
geographic information system (GIS) data can provide additional perspective and help in the
identification of contamination sources, potential sampling sites and nearby land uses. The
examination of historically accumulated data relating to microbiological results, beach postings
and disease surveillance will provide information on the area’s suitability for recreation and the
potential risks for swimmers. Assessment of hydrological, meteorological and other information
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Guidelines for Canadian Recreational Water Quality: Third Edition
on rainfall, currents, tides, prevailing winds and potential discharges (sewage, storm drains, other
waste discharges) can help identify their impact (either singly or collectively) on water quality.
2.2
On-site visit
The purpose of the on-site visit is to visually identify and confirm any and all existing or
potential water quality hazards. Information may similarly be collected on the existence and
adequacy of public facilities, safety provisions and mechanisms for public awareness and
communication. For the purposes of this EHSS, a hazard is an object or condition that may
endanger human health or safety. For most swimming areas, contact with faecal pollution in the
environment represents a significant concern; thus, attention should be paid to the potential
sources of faecal contamination, both point sources (discharge or drainage that may contain
sewage, stormwater or other faecal wastes) and diffuse sources (e.g., domestic and wild animals
and birds, stormwater runoff from the beach and surrounding areas, septic wastes, contamination
from swimmers themselves).
Additional existing or potential hazards can include:
• chemical hazards (e.g., industrial discharges, contamination from marinas/watercraft);
• biological hazards (e.g., cyanobacterial blooms, organisms responsible for swimmer’s itch);
and
• physical hazards (e.g., litter, poor visibility).
Other information collected may be useful in identifying hazards that are less visible. For
example, the presence of large amounts of floating debris may be indicative of sewage or
stormwater discharges. An example of an EHSS checklist, which includes the type of
information to be collected during an on-site visit, is provided in Appendix D.
Additionally, it is advisable to conduct site visits under both dry and wet weather conditions.
The effects of rainfall and storm events on water quality should be investigated. Certain
contamination events (e.g., runoff, stormwater discharges) may be visible only during rainfall
periods. Representative water samples may also be collected and analysed to confirm the
presence of contamination and determine its variability and source. Shortened surveys may also
be carried out throughout the swimming season at the time of microbiological monitoring to
collect more timely information about the recreational water area. Such information has
demonstrated value in developing models that may be capable of predicting water quality.
Further information on these topics can be found in Section 10.0 (Faecal pollution source
tracking) and Appendix B (Microbiological sampling and analysis).
2.3
Assessment report
Once the on-site visit is completed, a risk assessment should be performed to identify priority
water quality hazards. Risks, for the purpose of this EHSS, should consider the likelihood of
exposure to a given hazard and the associated consequence. Conducting a proper risk assessment
therefore requires the consideration of the factors that may contribute to a swimmer’s exposure.
These may include the proximity of the hazard to the swimming area, effects of the area’s
physical characteristics (depth, water circulation), potential weather influences, types and
patterns of recreational activities practised in the area and impacts of any existing barriers. For
example, in the case of the combined sewer overflow, factors contributing to swimmer exposure
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could be heavy rains causing a discharge of sewage material, currents or winds driving this
material towards the swimming area and the absence of public communication methods advising
that contact with the water should be avoided for a period immediately following heavy rainfall.
The risk assessment may also be used to identify potential points at which additional barriers
may be needed to reduce the degree of human exposure.
The process should culminate with the production of an assessment report, which should be used
when developing further beach management or operational plans. In addition to reporting on the
survey findings, the report should specify priorities for action, identify barriers that may be
implemented and provide recommendations for an appropriate beach monitoring program.
Recommandations for a monitoring program should identify specific sampling locations, times
and frequencies, as well as outline the steps to be taken in the event that a warning or other
actions are required.
The flowchart in Figure 1 (modified from Codd et al., 2005) is suggested as a possible sequence
of events when designing and implementing a multi-barrier strategy for recreational waters. It
may be used as a guide for beach operators, service providers or responsible authorities wishing
to develop their own operational plans.
EHSS:
Pre-Survey Preparations;
On-Site Visit
Situation Assessment
Assess Priorities for Action
Identify Potential Barriers
Appraisal of Feasible Barriers (Health,
Economic and Environmental Considerations)
EHSS:
Assessment Report
Selection of Option(s)
Formulate Option(s) into an Action Plan
Implement Action Plan
Monitor Effectiveness of Action Plan
Revise Plan as Necessary
Figure 1. Possible sequence of events for designing and implementing a multi-barrier strategy for recreational
waters.
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3.0
Barriers for protection of recreational waters
Barriers are procedures or actions that collectively reduce the risk of human exposure to
recreational water quality hazards. They can be physical actions, such as beach cleanup and
grooming, or they can be processes or tools that improve the effectiveness of the recreational
water management program, such as policies and legislation, guidelines and standards, and
education and communication strategies.
Four major areas have been described from which barriers can be identified (WHO, 2003a):
compliance monitoring; public awareness and communication; public health advice; and hazard
control actions.
3.1
Compliance monitoring
Monitoring is a broad concept and can serve many functions. It can be used to:
• determine whether water quality meets the Guidelines
• identify the impacts of water quality events;
• demonstrate long-term water quality trends;
• support EHSS findings or identify gaps;
• verify that barriers (e.g. notifications, corrective actions) are put in place;
• verify that these barriers are operating effectively.
Compliance monitoring is conducted to identify existing water quality hazards and to maintain a
record of changes that may occur. Proper monitoring and reporting are essential for assessing
and communicating information on the level of safety of recreational waters. Decisions regarding
the areas to be monitored, choice of indicators and monitoring program design will be made by
the appropriate regulatory and management authorities. The monitoring program should
incorporate information derived from the EHSS, taking into account recommendations made
regarding priority areas of concern. There should be a documented monitoring plan for all
monitored beaches, providing, at a minimum, instructions on:
• the parameters to be analysed;
• the locations at which samples are to be collected; and
• the times and frequencies of sample collection.
Recreational waters should be routinely monitored for the presence of faecal contamination,
using the following primary indicators of faecal contamination:
• Fresh water: E. coli;
• Marine water: Enterococci.
Guideline values have been established for the geometric mean and single-sample indicator
concentrations for these parameters (see Table 1 [Guidelines for Canadian recreational water
quality: summary table] and Section 4.0 [Recommended indicators of faecal contamination]).
Other organisms have also been described that may have potential value as indicators by
providing supplemental information with respect to the faecal contamination of recreational
waters. Such organisms may also be included as part of a recreational water monitoring program,
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Guidelines for Canadian Recreational Water Quality: Third Edition
provided they have been deemed appropriate by the responsible regulatory and management
authorities.
Other water quality parameters related to the physical or aesthetic characteristics of a
recreational water body and its surrounding environment have links to the health and safety of
recreational water users and thus may be included in a monitoring program. Guideline values or
aesthetic objectives have been specified for these parameters where they can be established (see
Table 1 [Guidelines for Canadian recreational water quality: summary table] and Section 8.0
[Physical, aesthetic and chemical characteristics]).
Tests may also be carried out for specific water quality hazards when there is epidemiological or
other evidence of their presence in water. Evidence may include:
• reports of a disease outbreak or illnesses of specific etiology;
• reports of a suspected illness of undetermined cause;
• reports of water-related injuries;
• levels of an indicator strongly suggesting the presence of a specific hazard;
• reports of a specific event such as a sewage or chemical spill; or
• the development of a cyanobacterial bloom.
Procedures may also be established for the monitoring of other barriers that may be utilized for
recreational water management. Examples may include:
• public notification / warning signs posted;
• water quality hazard control actions implemented;
• health authorities notified; and
• policies, legislation, protocols and/or guidelines established and in place.
It is recognized that recreational water areas will each have unique characteristics and
operational considerations. Decisions regarding the specific design of the recreational water
monitoring program should be made by the appropriate authorities (e.g. local, regional, federal,
provincial, territorial, as applicable). These should take into consideration the specific needs and
conditions of the area, the types of users and recreational activities practised, and any relevant
historical information.
3.1.1 Frequency of microbiological sampling
Decisions regarding the frequency of water samples collected for microbiological analysis
should be made by the appropriate local or regional authority. Guidance is provided on some of
the factors to be considered when selecting sampling frequency. Published texts are available
that can provide further information with respect to the design and implementation of
recreational water monitoring programs (e.g., Bartram and Rees, 2000).
There are a number of factors that can influence the microbiological quality of a recreational
water body at any given time. These can include the type and periodicity of contamination events
(both point and non-point sources), the time of day, recent weather conditions, the number of
users frequenting the swimming area and the physical characteristics of the area itself. The
significant day-to-day (and within-day) variation in indicator organism densities observed for
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recreational waters has been well documented (Leecaster and Weisberg, 2001; Boehm et al.,
2002; Whitman and Nevers, 2004; U.S. EPA, 2005a).
Results from the U.S. Environmental Protection Agency’s (EPA) Environmental Monitoring for
Public Access and Community Tracking (EMPACT) Project identified that it is the day-to-day
variation that is the principal source of uncertainty when attempting to develop an estimate of the
water quality for a recreational water area over a given period (U.S. EPA, 2005a). Correlations
were shown to exist only relative to indicator levels measured on the very next day. Indicator
levels demonstrated only a negligible correlation with those measured more than 2 days later.
More frequent monitoring (daily as opposed to weekly sampling; weekly as opposed to monthly
sampling) will have several advantages. As a result of the significant day-to-day variation in
faecal indicator counts that can be observed, even daily monitoring will not necessarily improve
the ability of the current day’s microbiological results to predict the next day’s water quality.
However, the additional information provided by increasing the number of samples will allow
the responsible authorities to more easily observe water quality trends and to make more
informed decisions regarding the area’s overall suitability for recreation. Moreover, it will enable
authorities to more quickly detect persistent water quality problems that may occur.
The Guidelines advocate the use of both a maximum limit for the geometric mean faecal
indicator concentration and a single-sample maximum limit. The use of dual limits allows
recreational water operators to better evaluate the water quality both in the short term and over
the duration of the swimming season. The single-sample limit will alert management to any
immediate water quality issues, whereas the geometric mean limit will alert management to
chronic contamination problems. This dual approach represents good monitoring practice as part
of an overall commitment to a strategy of risk management for recreational waters.
In order to ensure that human health is adequately protected, waters regularly used for primary
contact recreational activities should be monitored at a minimum frequency of once per week
during the swimming season. A weekly monitoring strategy is useful to alert managers and
responsible authorities to more persistent contamination problems that may have developed and
allows them to make the necessary decisions within a reasonable time frame. Weekly
surveillance is consistent with recommendations made by the U.S. EPA in its 2002
Implementation Guidance for Ambient Water Quality Criteria for Bacteria.
When sampling, consideration should also be given to the collection of samples for the purpose
of characterizing event-driven episodes of pollution that may affect recreational waters—for
example, immediately following periods of heavy rainfall or at times of greatest swimmer
activity.
In areas where high swimmer densities are expected, increased monitoring is recommended. In
such situations, the number of samples may be increased to permit the calculation of a weekly or
even daily geometric mean (based on a minimum of five samples), if so desired.
Similarly, certain circumstances may permit a reduction in the recommended sampling
frequency. These may include the existence of beaches in remote locations or in areas where
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Guidelines for Canadian Recreational Water Quality: Third Edition
primary contact recreational activities are not a regular occurrence or beaches that have
historically demonstrated acceptable water quality. Once an understanding of water quality
behaviour at a site has been achieved through relatively intensive monitoring and the use of an
Environmental Health and Safety Survey, a reduction in sampling frequency may be justifiable
and can help ease the burden of monitoring (Bartram and Rees, 2000; WHO, 2003a). Thus, if it
can be determined that a recreational water area is of consistently good microbiological quality,
does not have any obvious sources of faecal contamination and is not considered to present a
significant risk to the health and safety of its users, monitoring may be reduced to a frequency
sufficient to verify that the conditions have not deteriorated.
It may also be acceptable to reduce monitoring frequencies for recreational water areas that
consistently demonstrate poor water quality results, but only where appropriate management
actions are taken to discourage recreational use, and provided that the risks are clearly
communicated to the public.
3.1.2 Location of microbiological sampling
Decisions regarding the most appropriate location and depth of water samples collected for
microbiological analysis should be made by the appropriate local or regional authority.
Guidance is provided below on some of the factors to be considered when selecting sampling
locations and depths. Published texts are available that can provide further information with
respect to the design and implementation of recreational water monitoring programs (e.g.,
Bartram and Rees, 2000)
Most bodies of water used for recreational purposes are not completely homogeneous with
respect to their microbiological properties. In recreational water evaluations, the purpose of
sampling is to obtain aliquots that are as representative as possible of microbiological quality of
the area. A single water sample provides a quantitative estimate of the indicator bacteria present
at a particular site and time. Whitman and Nevers (2004) observed that there can be significant
variation between samples collected at multiple points along the beach, as well as among
samples collected within close proximity to each other at nearly the same time. As the total
number of samples increases, the more representative the data will be of the overall water
quality.
Sites should be chosen to be representative of the water quality encountered throughout the entire
swimming area. Consideration should be given to specific conditions that may influence the
levels and distribution of indicator organisms and pathogens. The sampling sites should include
points of greatest swimmer activity, as well as peripheral points subject to external faecal
pollution. Stormwater, sewage or river outlets can give certain sections of a body of water
microbiological qualities that are very different from those of the water body at large. The degree
of heterogeneity can also be affected by rainfall, wind direction and velocity, currents and tides,
or the presence of physical barriers, such as sandbars, natural or artificial wave breaks and piers.
The depth at which samples are collected can have a significant effect on the resulting estimates
of water quality (for the purposes of these Guidelines, ‘depth’ refers to the vertical distance in
the water extending from the bottom of the water to the surface). Where the water is very
shallow, disturbances of the foreshore sand and sediment caused by wave swash and swimmer
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activity can result in the resuspension of microorganisms. Where the water is deeper, this effect
has lesser influence on water quality measurements. In contrast, deeper waters are relatively
more exposed to offshore faecal contamination sources than shallower waters (U.S. EPA,
2005a).
Adult chest depth (approximately 1.2–1.5 m) has historically been the most common sampling
depth. Traditionally this has been considered to represent the depth of greatest swimmer activity
and the location nearest to the point of head immersion, which would be indicative of the risk
associated with accidentally swallowing water. Published epidemiological studies have typically
found that only samples collected at this depth show evidence of a mathematical relationship
between indicator organism density and swimmer illness.
Sampling at shallower depths (ankle or knee depth—approximately 0.15–0.5 m) may be more
representative of water quality encountered by young children playing at the water’s edge. It is
expected that more frequent swimming advisories would be issued if this monitoring approach
were used. Sand and sediment disturbances can result in increased microbiological numbers in
shallower waters. Currently there is insufficient evidence to determine whether the expected
increase in the number of swimming advisories at this monitoring depth would result in a
proportionate reduction in the number of swimmer illnesses
Another strategy for monitoring that has been proposed involves the attempt to strike a balance
between the depth at which the majority of the health effects have been proven and the depth at
which microbiological counts are thought to be the highest (U.S. EPA, 2005a). According to the
recommendations outlined in the U.S. EPA’s EMPACT report (U.S. EPA, 2005a), sampling in
water of knee to waist depth may offer a reasonable, but still conservative, approach to
monitoring.
Alternatively, another approach would be for authorities to sample at multiple depths—for
example, at ankle to knee depth, as well as at waist to chest depth. Such a design could be used
to produce separate estimates of the water quality in both shallow waters and waters of
swimming depth. However, when comparing the results of water samples between shallow and
deeper waters, samples collected at a specific depth should be analysed as a singular group in
order to improve the precision of the data.
The sample depth (or depths) selected for an individual beach should be determined by the local
or regional authority in order to obtain the best information for their particular recreational water
area.
3.1.3
Other monitoring tools
Composite sampling
The use of composite sampling techniques presents a possible means for increasing the area
covered under a beach monitoring program, while potentially minimizing the costs associated
with analysis. Composite sampling involves the collection of multiple samples from across a
stretch of beach, combining them into one large composite, and then analysing a subsample of
the resulting mixture.
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There are obstacles to be overcome with the use of this technique. Increased sampling is required
initially to validate whether composite sampling will be feasible at a given area. As well the area
must be characterized to identify hot spots (sampling points having continuously poor water
quality) that can disrupt the analysis. Some statistical knowledge is also required to analyse the
data. Nevertheless, preliminary investigations have suggested that, if properly conducted,
composite sampling can be used in making water quality decisions with a comparable degree
of accuracy to that achieved by analysing the samples individually and averaging the results
(Kinzelman et al., 2006). Further information on composite sampling can be found in Appendix
C (Composite sampling for faecal contamination).
Predictive water quality models
An emerging area of research has been the development of predictive models that are able to
make same-day predictions about the microbiological quality of the water. Researchers have
developed and validated models for specific beaches to make predictions of water E. coli levels
using data from various water- and weather-related parameters (e.g., rainfall, wave height, wind
direction, turbidity, previous day’s faecal indicator counts). There are a few models currently
used in beach monitoring programs in the United States:
• SwimCast, in use at several Lake Michigan beaches in Lake County, Michigan uses air and
water temperature, wind speed and direction, precipitation, relative humidity, wave height,
lake stage, sunlight and other water quality parameters to predict whether current E. coli
levels are suitable for swimming (Olyphant and Pfister, 2005; Lake County Health
Department, 2010).
• Project S.A.F.E. (Swimming Advisory Forecast Estimate) in use at five Lake Michigan
beaches in northern Indiana uses wind direction, gauge height of the nearby Burns Ditch,
rainfall in the previous 48 hours and lake conditions such as chlorophyll and turbidity to
make predictions about relative E. coli levels and swimmability (Whitman, 2005; U.S.
Geological Survey, 2007).
• Nowcasting in use at Huntington and Edgewater beaches on Lake Erie in Ohio makes use of
log10 turbidity data, wave height and radar rainfall from grids surrounding the respective
beach areas during the previous 24 or 48 hours to estimate if E. coli levels are acceptable for
swimming at these beaches (Francy, 2007).
Evidence generated to date suggests that a properly developed model can achieve a degree of
accuracy comparable to that achieved by traditional approaches that use the previous day’s
indicator concentrations.
There are a number of challenges associated with model development. A significant level of
technical expertise is required to develop the models and to analyse the data, and models may not
work in all areas. Nevertheless, they do present a possible means for forecasting the quality of
the water, thereby improving the timeliness of decisions of whether to open or close a beach.
Beach operators, service providers or responsible authorities looking for an additional tool with
which to potentially improve the timeliness of their water quality decisions may wish to
investigate this approach.
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3.2
Public awareness and communication
In order to participate in safe, enjoyable recreational water activities, the public requires access
to information on the quality of the area and its facilities, as well as notification of any existing
water quality hazards. Beach operators, service providers and responsible authorities have a
responsibility to inform and educate the public and provide adequate warnings about any hazards
relevant to their recreational water areas.
Efforts to improve the public’s awareness and understanding of water quality can have numerous
benefits (Bartram and Rees, 2000; Pendleton et al., 2001). Communication tools can be used to:
• reduce the potential risk of swimmer illness or injury;
• improve the quality of the water;
• correct public misconceptions regarding water quality;
• improve public confidence; and
• increase beach attendance.
Posting of information at recreational water areas
Information on the quality of the water should be communicated to the user through the use of
posted signs. Signs should be used to warn users when the water is unsafe for recreational use as
well as to communicate when the water is safe for use.
Signs should be posted in locations that are highly visible to the public. The information
provided should be easily understood and not open to misinterpretation. Ideally, signs should be
standardized to permit comparisons across different locations. Warnings should be timely and
should be promptly removed once the issuing authority has determined that the risk no longer
exists.
The information provided on warning signs should include, at a minimum:
• a statement identifying the health or safety risk;
• recommended actions to be taken;
• the name of the authority responsible for issuing the warning; and
• contact information for the issuing authority.
Similarly, during periods in which the area is considered suitable for use, corresponding signs
should be posted that clearly communicate this information to the public. One important concept
to communicate to beach users through education is that even in waters considered of good
quality for swimming, there is always some probability that swimmers may experience some
adverse health effects.
Examples of informative beach signs are provided in Appendix F.
There are two main situations under which a warning sign may be posted—following the issuing
of a swimming advisory and following the issuing of a beach closure. Issuing a swimming
advisory or a beach closure should be made by the Medical Officer of Health or other
appropriate authority in accordance with the statutes existing in each province or territory. This
decision should be based on a thorough assessment of the situation with information provided by
recreational water monitoring, the EHSS and existing public health surveillance.
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A swimming advisory can be issued if the responsible authority identifies that the water is not
suitable for recreational use. Under this situation, users are advised to refrain from whole body
contact with the water. Contact with the beach is usually permissible, and access to the facilities
is generally not restricted. Examples of scenarios that may trigger jurisdictions to decide to issue
a swimming advisory include:
• exceedance of the guideline values for the recommended indicators of faecal contamination;
• exceedance of the guideline values for toxic cyanobacteria and their toxins, or in the event of
the development of a cyanobacterial bloom;
• existence of evidence of the risk of swimmer’s itch for recreational water users; and
• after periods of significant rainfall, which could trigger an advisory as a pre-emptive action.
A beach closure can be issued if the responsible authority identifies that a beach or body of water
poses a serious risk to the health and safety of recreational water users, and that it is further
necessary to restrict individuals from coming in contact with the area. Under a closure, the area
is considered closed to all recreational activity. Users are advised to avoid contact with the beach
and recreational water area, and access to the facilities may be restricted. Examples of situations
in which jurisdictions may determine it necessary to warrant the issuing of a beach closure
include:
• suspicion that the area is responsible for a waterborne disease outbreak;
• a sewage or chemical spill that is expected to affect the recreational water area;
• other conditions such that the area is judged to pose a significant risk to public health (e.g.,
persistently poor water quality); and
• detection of a cyanobacterial bloom.
Appropriate signs may also be posted at waters that have been deemed suitable for secondary
contact activities (e.g., rowing, sailing, canoe touring, fishing), but not for primary contact uses
(e.g., swimming, wading, windsurfing, waterskiing). In these instances, it may be necessary to
expand the location of the signs beyond the beach area to improve their visibility. Suggested
locations include relevant points of entry and launch areas.
The public can also do their part by: educating themselves on actions they can take to protect
themselves and the beach; becoming aware of where the water quality monitoring results are
posted; and, consulting this information before going to the beach.
Other tools for public education and communication
In addition to posted signs, other tools can be used for public education and the communication
of information, including:
• printed materials (e.g., posters, information sheets, educational bulletins, pamphlets,
brochures);
• media sources (e.g., local newspapers, television and radio announcements, Internet
websites);
• participation in beach certification or award programs;
• educational events such as volunteer monitoring programs and beach cleanup days; and
• classification or grading systems for beaches.
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Classification or grading systems for beaches have received interest as a tool to promote
communication and understanding of water quality information. They are also thought to help
encourage a sense of shared accountability and responsibility among the beach authorities and
beach users. A number of jurisdictions and multinational organizations have adopted grading
systems as part of their recreational water management recommendations (WHO, 2003a;
MDDEP, 2004; NHMRC, 2008). Both the WHO and the Australian Guidelines (WHO, 2003a;
NHMRC, 2008) make use of a grade-based format for faecal indicator density within their
framework for assessing faecal pollution in recreational waters. These approaches use a
microbiological assessment component along with a sanitary inspection categorization to
produce a classification of an area’s overall suitability for recreation. Under the Quebec
framework for faecal pollution assessment (MDDEP, 2004), water quality grades are assigned to
recreational waters based on average yearly faecal indicator monitoring results. This grade is
used as a public communication tool and also dictates the sampling frequency requirements for
the area.
There are advantages and disadvantages associated with the use of beach grading systems.
Responsible authorities should be aware of the limitations of any system when investigating its
use as a potential communication tool. For example, grading systems based on faecal indicator
monitoring results provide information on only one aspect of recreational water safety.
Moreover, the accuracy of any such grading system would be strongly affected by the limitations
known to be associated with faecal indicator monitoring. Ideally, a successful beach grading
system would involve criteria from a number of categories, capturing monitoring results,
communication tools and water quality hazard control actions in place.
3.3
Public health advice
Consultation with public health authorities is another essential component of risk management.
In the event of an incident (microbiological, chemical or physical) that represents a risk to public
health or safety, health officials can play a key role by providing advice and determining what
actions need to be taken. Local public health authorities should be promptly notified of any
situation that threatens the health or safety of recreational water users. Similarly, as part of
normal operations, local public health officials may be periodically consulted for information
and advice on topics pertinent to safe recreational water use.
In assessing the risks associated with recreational water quality hazards, the local health
authorities should, wherever possible, establish surveillance for swimmer illness or injuries. This
can be established by consulting public health surveillance mechanisms currently in operation or
by conducting specific investigations. Information sources include:
• federal, provincial/territorial or regional departments or agencies having surveillance
programs or reporting systems;
• clinical reports from hospital emergency departments and local physicians;
• accident or incident reports held by recreational water area operators or service providers;
• formal epidemiological investigations; and
• other potential surveillance mechanisms (e.g., monitoring of over-the-counter medicine sales
in pharmacies).
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Guidelines for Canadian Recreational Water Quality: Third Edition
Procedures for the investigation of illness associated with recreational waters should adhere to
the recommendations given in the second edition of Procedures to Investigate Waterborne
Illness (International Association for Food Protection, 2002).
3.4
Hazard control actions
Hazard control actions are physical actions intended to reduce the impact of microbiological,
chemical or physical water quality hazards on a particular recreational water site. A discussion
on the numerous types of possible hazard control actions is outside the scope of this document.
In addition, the types of hazard control actions required and their relative effectiveness will be
specific to each beach and each situation. Authorities may wish to consult the following types of
resources for information to help them address specific beach-related issues:
•
•
•
•
Published texts: Texts are available which provide comprehensive discussion of larger topics
like stormwater management, wastewater treatment and coastal water management.
Searchable databases of journal articles: Keyword searches can direct the user to specific
scientific studies, related articles, bibliographical citations and topic review papers.
Proceedings from international conferences: Review of proceedings can identify accounts of
actions evaluated in other communities, providing an indication of results and potential
contacts; specific conferences include: Great Lakes Beach Association Conferences
(www.great-lakes.net/glba), and the U.S. EPA National Beach Conferences (www.epa.gov).
Manuals or publications produced by stakeholder organizations (e.g., Griffiths, 1999).
Assessment may help identify smaller-scale actions that may provide good returns from a cost–
benefit perspective. However, some issues may be more substantial in nature and thus may not
be easily resolved without the application of more sophisticated control methods. Examples of
small-scale control actions aimed at reducing faecal contamination can include beach cleanup or
grooming procedures or installation of structures (fences, overhead wires) to discourage the
presence of birds and other wildlife. Examples of larger-scale actions can include sewage
treatment or storm drain waste retention. Any potential control action should be assessed from
the perspectives of the effects on the health of the users and on the environment.
Water quality issues can cross over multiple boundaries (e.g., health, environment, agriculture,
municipal infrastructure), and require cross-sectoral collaboration. Consultation with responsible
authorities, other beach operators or service providers and recreational water quality
professionals may help to identify actions that have proven to be successful in other
communities.
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Guidelines for Canadian Recreational Water Quality: Third Edition
Part II: Guideline Technical Documentation
4.0
Recommended indicators of faecal contamination
Recreational waters may be contaminated with faecal material from such sources as discharged
sewage, stormwater runoff from agricultural or urban areas, wild or domesticated animals, and
even through faecal shedding by swimmers themselves. Many epidemiological studies have
identified gastrointestinal and upper respiratory illnesses in swimmers as a result of such
contamination. Historically, the bacteria in the coliform group and its subgroups (total coliforms,
thermotolerant [faecal] coliforms, E. coli) and the enterococci—the more faecal-specific portion
of the faecal streptococci group—have been used to monitor recreational waters for the presence
of faecal contamination. As such, they have also been used to indicate the possible presence of
pathogenic microorganisms responsible for these illnesses. Routine testing of recreational waters
for pathogenic organisms is impractical and is not recommended for the following reasons:
• Testing for every possible waterborne disease-causing microorganism would be prohibitive
in terms of both the financial resources necessary and the time required to perform the
analyses. These organisms are difficult to isolate and quantify, and testing requires proper
laboratory containment facilities, specialized equipment and highly trained and experienced
microbiologists. Detection methods for some pathogens do not exist at all.
• Pathogens are usually present at low levels and are unevenly distributed in recreational
waters, even during disease outbreaks.
• The absence of one pathogen does not necessarily ensure that other enteric pathogens are
also absent.
Consequently, authorities monitor for non-pathogenic faecal indicator bacteria that are present in
high numbers in both human and animal faeces. Elevated numbers of these indicator bacteria in
the aquatic environment are indicative of faecal contamination and therefore suggest the possible
presence of enteric pathogens.
The ideal faecal indicator organism would meet the following requirements (Cabelli et al., 1983;
Elliot and Colwell, 1985):
• found within the intestinal tract of humans and warm-blooded animals;
• present in faecally contaminated waters when enteric pathogens are present, but found in
greater numbers than pathogens;
• incapable of growth in the aquatic environment, but capable of surviving longer than
pathogens;
• applicable to all types of natural recreational waters (fresh, estuarine and marine waters); and
• absent from non-polluted waters and exclusively associated with animal and human faeces.
Other desirable qualities for the indicator organism include:
• Density of the indicator should be directly correlated with the degree of faecal contamination.
• Density of the indicator should be quantitatively related to swimmer-associated illnesses.
• Detection and enumeration test methods should be rapid, easy to perform, inexpensive,
specific and sensitive.
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Guidelines for Canadian Recreational Water Quality: Third Edition
No single microorganism unequivocally meets all of these criteria. E. coli and enterococci are
currently considered the best indicators of faecal contamination in recreational waters, as they
most closely fit the above characteristics. There are limitations associated with the use of
indicators in assessing the quality of recreational waters. Judicious use of these guideline values
as part of a multi-barrier approach to recreational water management represents a sound
approach to protecting swimmers against exposure to faecal pathogens in the recreational water
environment.
4.1
Indicator organisms for primary contact recreation
4.1.1
Fresh waters: Escherichia coli (E. coli)
Guideline values
For fresh recreational waters used for primary contact activities, the guideline values are as
follows:
Geometric mean concentration (minimum of five samples):
≤ 200 E. coli/100 mL
Single-sample maximum concentration:
≤ 400 E. coli/100 mL
Calculation of the geometric mean concentration should be based on a minimum of five samples,
collected at times and sites so as to provide representative information on the water quality likely
to be encountered by users. Further action should be initiated if either of these guideline values is
exceeded. Minimum action should consist of immediate resampling of the site(s). In addition, a
swimming advisory may be issued should the responsible authority identify that the area is not
suitable for recreational water use.
It is further advised that recreational water areas routinely used for primary contact recreation be
monitored at a minimum of once per week, with increased monitoring recommended for those
beaches that are highly frequented or are known to experience high user densities. Similarly,
under certain scenarios, a reduction in the recommended sampling frequency may be justified.
Further guidance on sampling frequency recommendations and the posting of recreational waters
can be found in Part I (Management of Recreational Waters).
Enterococci (Section 4.1.2) is also recognized as a suitable indicator of faecal contamination in
fresh recreational waters (Cabelli, 1983; Pruss, 1998; Wade et al., 2003, 2006). If it can be
shown that enterococci can adequately demonstrate the presence of faecal contamination in fresh
waters, then the enterococci maximum limits for marine waters may be adopted. If there is any
doubt, samples should be examined for both sets of indicators for extended periods to determine
whether a positive relationship exists.
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Guidelines for Canadian Recreational Water Quality: Third Edition
Guideline rationale
The guideline values have been developed based on epidemiological evidence relating
E. coli concentrations in fresh recreational waters to the incidence of swimming-associated
gastrointestinal illness observed among swimmers. The existing epidemiological data are not
sufficient to permit the estimation of the level of risk for individual exposures. Based on the U.S.
EPA’s regression analysis of epidemiological data (Dufour, 1984), Health Canada has estimated
that using the guideline values for the recommended indicators of faecal contamination for fresh
and marine waters will correspond to a seasonal gastrointestinal illness rate of approximately 1–
2% (10–20 illnesses per 1000 swimmers).
In determining the value for the maximum faecal indicator concentration permitted in a single
sample, the U.S. EPA equations pertaining to the calculation of a single-sample limit were
reviewed (U.S. EPA, 1986). The data regarding the maximum allowable indicator density at
designated beach areas is consistent with applying a factor of 2 times the recommended
geometric mean value results. Thus, the single-sample maximum concentration of 400 E.
coli/100 mL is reaffirmed.
These values represent risk management decisions that have been based on a thorough
assessment of the potential risks for the recreational water user. In considering both the potential
health risks and the benefits of recreational water use in terms of physical activity and
enjoyment, it was concluded that this is a tolerable and reasonable estimate of the risk of illness
likely to be experienced by users engaged in a voluntary activity.
An evaluation of the epidemiological information published since the Guidelines were last issued
concluded that the current body of evidence supports the existing recommendations regarding the
use of E. coli as the indicator of faecal contamination in fresh recreational waters. There has not
been substantial evidence to suggest that revision of the existing guideline values is necessary at
this time.
Description
E. coli most closely fits the requirements of an ideal indicator of faecal contamination for fresh
waters. The organisms are found in high numbers in the intestinal tract and faeces of humans and
warm-blooded animals. The vast majority of the E. coli types are harmless. There are several
types (serotypes or strains) that possess virulence factors enabling them to act as human
pathogens; however it should be noted that faecal concentrations of the typical non-pathogenic E.
coli will always be greater than those of the pathogenic strains, even during outbreaks. E. coli is
considered a more specific indicator of faecal contamination than either total coliforms or
thermotolerant (faecal) coliforms and can be rapidly and easily enumerated in recreational
waters.. In addition, a strong correlation has been demonstrated between the concentration of E.
coli in fresh waters and the risk of gastrointestinal illness among swimmers (Dufour, 1984; Wade
et al., 2003).
For several decades now, recreational water quality experts in Canada have recognized E. coli as
the indicator of choice for faecal contamination. The use of E. coli as an indicator of recreational
water quality was of limited use until the 1980s, when standardized laboratory methods
permitting detection of the organism within 24–48 hours became available. Prior to this, the
27
Guidelines for Canadian Recreational Water Quality: Third Edition
thermotolerant coliform group was used as the primary indicator of faecal contamination in
recreational waters. However, subsequent findings that some thermotolerant coliform species had
non-faecal or environmental origins and could be isolated in high numbers from waters receiving
waste effluents from such sources as pulp and paper and textile manufacturing facilities (Dufour
and Cabelli, 1976; Huntley et al., 1976; Rokosh et al., 1977; Vlassoff, 1977) raised concerns
about the reliability of using this group as an indicator of faecal contamination in recreational
waters. Despite the availability of methods specific for the detection of E. coli, testing
laboratories were already set up to test for thermotolerant coliforms, and the requirements to
perform surveillance of recreational waters for these microorganisms were embedded in longstanding regulatory and legislative documents. As a result, it has taken many years and
considerable time to amend the existing guidelines and standards, updating them to reflect the
current state of knowledge that E. coli is the preferred indicator of faecal pollution for fresh
recreational waters.
In the 1992 edition of the Guidelines, the introduction of E. coli as the recommended indicator
for freshwater quality represented a new direction for recreational water monitoring, moving
away from the 1983 recommendation regarding the use of thermotolerant coliforms. As a result,
a provision was made that allowed thermotolerant coliforms to continue to be used if it could be
determined that greater than 90% of the thermotolerant coliforms were, in fact, E. coli. This was
done to permit an adjustment period for jurisdictions in making the changeover to the new
recommendations. Significant time has now passed to allow jurisdictions to make the change
from thermotolerant coliforms to the more faecal-specific E. coli. As a result, this third edition
does not recommend the use of thermotolerant coliforms as an indicator of the quality of
recreational waters. Instead, it reaffirms that E. coli is the preferred indicator for monitoring
fresh recreational waters in Canada.
Occurrence in the aquatic environment
Within human and animal faeces, E. coli is present at a concentration of approximately 109 cells
per gram (Edberg et al., 2000) and comprises about 1% of the total biomass in the large intestine
(Leclerc et al., 2001; Health Canada, 2012a). Human faecal flora characterization studies have
reported that E. coli was detected in 94% and 100% of the subjects tested (Finegold et al., 1983;
Leclerc et al., 2001). These values were significantly higher than those reported for other
members of the coliform group and were matched or exceeded by only enterococci and certain
species of anaerobic bacteria (Bacteroides, Eubacterium).
E. coli comprises about 97% of the coliform organisms in human faeces, with Klebsiella spp.
comprising 1.5% and Enterobacter and Citrobacter spp. together comprising another 1.7%.
E. coli has been shown to represent between 90% and 100% of all coliforms in faeces from
eight species of domestic animals, including chickens (Dufour, 1977).
Once shed from a human/animal host, faecal bacteria are not expected to survive for long periods
in the aquatic environment (Winfield and Groisman, 2003). Survival of E. coli in the recreational
water environment is dependent on many factors, including temperature, exposure to sunlight,
available nutrients, water conditions such as pH and salinity, and competition from and predation
by other microorganisms.
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Guidelines for Canadian Recreational Water Quality: Third Edition
Numerous authors have reported on the ability of beach sand and sediments to prolong the
survival of faecal microorganisms (Whitman and Nevers, 2003; Ishii et al., 2006a; Kon et al.,
2007a). This environment is thought to provide more favourable conditions of temperature and
nutrients than the adjacent waters, as well as to offer protection from certain environmental
stressors such as sunlight. Others have reported on the ability of E. coli to survive in organic-rich
environments not known to be associated with faecal contamination, such as industrial process
wastes and wastes from pulp and paper manufacturing (Megraw and Farkas, 1993; Gauthier and
Archibald, 2001). Recently, researchers have reported on the ability of E. coli and other faecal
bacteria to survive within mats of the green algal species Cladophora (Whitman et al., 2003;
Olapade et al., 2006).
Historically it was believed that E. coli could only originate from faecal sources and was
incapable of growth in the aquatic environment. However, recent studies are raising questions
regarding these assumptions (Kon et al., 2007b; Hartz et al., 2008, Vanden Heuvel et al., 2009).
Research in this area is ongoing. These recent findings do not invalidate the use of E. coli as the
best available indicator for recreational water quality.
E. coli is considered to be a good surrogate of the survival of enteric bacterial pathogens in
recreational waters. Several authors have reported similar survival rates for E. coli and enteric
bacterial pathogens (Rhodes and Kator, 1988; Korhonen and Martikainen, 1991; Chandran and
Mohamed Hatha, 2005). The indicator is regarded to be more sensitive to environmental stresses
than human enteric viruses and protozoa and thus does not survive as long in the environment as
these organisms.
In many parts of Canada, freshwater beaches are routinely monitored for levels of E. coli as
an indication of faecal contamination. Many Canadian recreational waters are of good
microbiological quality; however, certain waters are contaminated throughout part, or all, of the
swimming season. An examination of beach monitoring data over a 10-year period (1993–2003)
at 10 Lake Huron recreational beaches in Ontario demonstrated that levels of E. coli can vary
widely at a single location from year to year and between beach locations (Ontario Ministry of
the Environment, 2005). E. coli values can range from 0/100 mL in isolated areas to several
thousand per 100 mL in areas directly impacted by faecal contamination (Payment et al., 1982;
Sekla et al., 1987; Williamson, 1988; Ontario Ministry of the Environment, 2005).
Association with pathogens
E. coli is considered a good indicator for enteric bacterial pathogens such as Salmonella,
Shigella, Campylobacter and E. coli O157:H7 (Health Canada, 2012a). Investigations conducted
by the Water Environment Research Foundation (Yanko et al., 2004) examined the relationship
between E. coli concentrations in surface water samples collected from various watersheds in
southern California and the probability of detecting Salmonella and Shiga toxin-producing
E. coli (STEC). The results demonstrated that the probability of detecting Salmonella using
culture-based methods steadily increased up to a concentration of approximately
1000 E. coli/100 mL. At this point, a 100% probability of detection was reported. Similar
results were reported for the detection of STEC strains using polymerase chain reaction (PCR)based methods. Although there was clearly an association between E. coli concentrations and
the probability of detection of Salmonella and STEC strains, the authors reported that no single
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Guidelines for Canadian Recreational Water Quality: Third Edition
sample could be used to provide absolute assurance of the presence or absence of these
pathogens.
E. coli is a less effective indicator of enteric pathogenic viruses and protozoa. Numerous studies
have reported on the lack of a correlation between E. coli concentrations and the presence of
enteric viruses and protozoa in surface waters (Griffin et al., 1999; Denis-Mize et al., 2004;
Hörman et al., 2004; Dorner et al., 2007).
E. coli are always present in faecal contamination from human and animal sources. Detection
indicates faecal contamination of water and thus the possible presence of faecal pathogenic
bacteria, viruses and protozoa. The occurrence of faecal pathogens in recreational waters is
strongly dependent on the nature of the contamination sources impacting the swimming area.
Their presence and numbers in the environment can be sporadic and highly variable. As well,
some enteric pathogens can survive longer than the faecal indicators. The absence of E. coli
should not be interpreted to mean that enteric pathogenic microorganisms are also absent.
The combination of routine E. coli monitoring alongside actions, procedures and tools to
collectively reduce the risk of swimmer exposure to faecal contamination in the recreational
water environment represents the most effective approach to protecting the health of recreational
water users.
Guidelines used by other countries/organizations
The guideline formats and values established by other government and multinational
organizations worldwide for faecal indicator organisms in fresh waters were reviewed in
developing the revised edition of this document.
Table 2. Guideline values for faecal indicator concentrations in fresh recreational waters established by other
countries or organizations.
Country/
Freshwater
Format and guideline values
Reference
organization
indicator
U.S EPAa
E. coli
Geometric mean concentration:
U.S EPA, 2002
126/100 mL
Single sample maximum concentration :b
235/100 mL
Enterococci
Geometric mean concentration:
33/100 mL
Single sample maximum concentration:b
62/100 mL
WHO
Intestinal
95th percentile/100 mL:
WHO, 2003a
enterococcic
A: ≤40
B: 41-200
C: 201-500
D: >500
Australia
Intestinal
95th percentile/100 mL:
NHMRC, 2008
enterococcic
A: ≤40
B: 41-200
C: 201-500
D: >500
European Union
Intestinal
95th percentile/100 mL:
EU, 2006
enterococci
Excellent: 200 /100 mL
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Guidelines for Canadian Recreational Water Quality: Third Edition
Good: 400/100 mL
E. coli
90th percentile/100 mL:
Sufficient: 330/100 mL
95th percentile/100 mL:
Excellent: 500 /100 mL
Good 1000/100 mL
90th percentile/100 mL:
Sufficient: 900/100 mL
a – Note that new criteria are currently under development and may be available in 2012.
b – Designated beach area (75% confidence level)
c – Recommends guidelines for coastal waters be used until more freshwater data is available.
Related epidemiological studies
The U.S. EPA’s original epidemiological study in fresh recreational waters.
measured the concentrations of faecal indicator organisms (thermotolerant (faecal) coliforms, E.
coli, enterococci) in swimming waters and compared them with the rates of swimmingassociated illness reported on the same days on which the samples were collected (Dufour,
1984). Statistically significant rates of gastrointestinal illness were observed among swimmers in
waters considered to be more faecally contaminated. For symptoms unrelated to gastrointestinal
illness, no statistically significant differences were observed. For the data analysis, the mean
seasonal faecal indicator concentration per 100 mL was plotted against the seasonal swimming
associated rate of gastrointestinal symptoms per 1000 persons for each indicator. Correlation
and regression analysis were then used to determine the correlation coefficients and the slope of
the linear regression equation for each indicator. The best correlation coefficient (r) was
obtained with E. coli (r = 0.80) and an almost equal correlation coefficient was observed with
enterococci (r = 0.74). The E. coli data were used to produce a regression equation:
seasonal risk of gastrointestinal illness
= 9.40 (log E. coli concentration per 100 mL) – 11.74
per 1000 persons
Several freshwater epidemiological studies have been conducted since the previous version of
the Guidelines for Canadian Recreational Water Quality was developed (Lightfoot, 1988;
Ferley et al., 1989; Calderon et al., 1991; van Asperen et al., 1998). All have confirmed the
existence of a strong relationship between exposure to recreational waters and swimmingassociated illness; however, few have been able to demonstrate evidence of a mathematical
relationship between faecal indicator counts and illness among swimmers. Van Asperen et al.
(1998) reported that the risk of gastroenteritis was significantly higher among triathletes who
swam in water with a geometric mean concentration of > 355 E. coli colony-forming units
(cfu)/100 mL (equivalent to E. coli/100 mL). Ferley et al. (1989) proposed that faecal
streptococci were a better indicator of gastrointestinal illness at freshwater beaches in France.
Calderon et al. (1991) observed that total staphylococci counts were most closely associated with
gastrointestinal illness among swimmers at a recreational pond not affected by point source
discharges.
Only a few epidemiological studies have been conducted that have investigated the health effects
of recreational activities other than swimming, such as whitewater canoeing and rafting (Fewtrell
et al., 1992; Lee et al., 1997). The data linking water quality and illness through these activities
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Guidelines for Canadian Recreational Water Quality: Third Edition
have been less strong. Still, conclusions that have been reported are that gastrointestinal illness
similarly constitutes the most frequently reported health outcome during these types of activities;
and that factors related to the risk of illness include the quality of the water and the frequencies
of immersion and water ingestion.
Several reviews of the epidemiological findings have also been published. In 1998, WHO (Pruss,
1998) published a comprehensive review of the epidemiological research that had been
conducted over the period from 1953 to 1996. This was the first extensive review of the existing
epidemiological literature. Pruss (1998) concluded that gastrointestinal illness was the most
frequent health outcome for which dose–response relationships were reported and that the
indicators that best correlated with health outcomes were enterococci for marine waters and
E. coli and enterococci for fresh waters.
The U.S. EPA published two reviews of the existing epidemiological literature for recreational
waters. The first, published in its Implementation Guidance for Ambient Water Quality Criteria
for Bacteria (U.S. EPA, 2002), was a minireview of the epidemiological studies conducted since
the previous guidance had been issued, in 1986. The U.S. EPA concluded that the
epidemiological methods used to derive its 1986 Water Quality Criteria remained scientifically
valid and that no new scientific principles had been established that would require a revision of
the current guidelines. Subsequent to this, the U.S. EPA (Wade et al., 2003) conducted a metaanalysis of the existing epidemiological data available in the literature to determine if its current
regulatory standards were sufficiently protective against the risk of gastrointestinal illness in
recreational waters. The authors demonstrated that in the freshwater studies, E. coli was shown
to be the most suitable indicator of recreational water illness. It was further reported that when
comparing the summary relative risk values with the U.S. EPA guidelines for fresh water, E. coli
densities above the guideline values were associated with an increased risk of illness, whereas
exposures below the guideline values did not show an association.
Wiedenmann et al. (2006) reported on the results of a randomized controlled prospective cohort
study conducted at freshwater swimming sites in Germany. Earlier randomized controlled trials
had been conducted in coastal waters of the United Kingdom (Kay et al., 1994; Fleisher et al.,
1996); however, this study was the first of its type to be conducted in fresh waters. The study
design used by the authors was similar to that originally used in the UK trials. The authors
observed evidence of a relationship between the observed rates of illness and measured
concentrations of E. coli, enterococci, Clostridium perfringens and somatic coliphages. Noobserved-adverse-effect-levels (NOAELs) were reported for several definitions of
gastroenteritis, ranging from 78 to 180 E. coli/100 mL and from 21 to 24 enterococci/100 mL.
The authors proposed possible guidelines by combining all of the data derived from the different
definitions of gastrointestinal illness investigated, suggesting values of 100 E. coli/100 mL, 25
enterococci/100 mL, 10 somatic coliphages/100 mL and 10 C. perfringens/100 mL. Although the
authors propose a value of 100 E.coli/100 mL, it is important to note that the NOAEL reported
for gastrointestinal illness that most closely fits the criteria of “highly credible gastrointestinal
illness” (as defined by Cabelli et al , 1983) was 180 E. coli/100 mL. In addition, the quartile and
quintile breakdown of the data for the UK definition of gastrointestinal illness indicated that the
rates of swimmer illness compared to those of the control group were not statistically significant
until E. coli concentration ranges approached or exceeded 200 E. coli/100 mL. Even using the
32
Guidelines for Canadian Recreational Water Quality: Third Edition
least stringent definition of gastrointestinal illness resulted in a NOAEL (NL-2 1: 164 E. coli/100
mL) well above 100 E. coli/100 mL.
The U.S. EPA and Centers for Disease Control and Prevention (CDC) have also been conducting
epidemiological studies at freshwater and marine beaches under the National Epidemiologic and
Environmental Assessment of Recreational (NEEAR) Water Study. The studies are intended to
support the development of new recreational water quality guidelines (U.S. EPA, 2002; Dufour
et al., 2003) as well as to investigate new water quality indicators and rapid methods for water
quality monitoring. Data collection for these studies was completed in 2010.
Summary
It has been concluded, based on all of the existing evidence, that E. coli remains the most
suitable indicator of faecal contamination in fresh recreational waters. In summary:
1. The guideline values have been developed based on the analysis of epidemiological evidence
relating E. coli concentrations in fresh recreational waters to the incidence of swimmingassociated gastrointestinal illness observed among swimmers. The values represent risk
management decisions based on the assessment of possible health risks for the recreational
water user and the recognition of the significant benefits that recreational water activities
provide in terms of health and enjoyment.
2. E. coli most closely fits the requirements of an ideal indicator of faecal contamination for
fresh waters. E. coli are always present in faecal contamination from human and animal
sources. Detection suggests faecal contamination of water and thus the possible presence of
faecal pathogenic bacteria, viruses and protozoa.
3. There are limitations associated with the use of indicators in assessing the quality of
recreational waters. The occurrence of faecal pathogens in recreational waters is dependent
on many factors and can be variable and sporadic. The absence of E. coli should not be
interpreted to mean that enteric pathogenic microorganisms are also absent.
4. Combining routine E. coli monitoring alongside actions, procedures and tools to collectively
reduce the risk of swimmer exposure to faecal contamination in the recreational water
environment represents the most effective approach to protecting the health of recreational
water users.
1
NL-2 – Definition of gastrointestinal illness used in the Netherlands according to Van Asperen et al., 1998.
33
Guidelines for Canadian Recreational Water Quality: Third Edition
4.1.2
Marine waters: Enterococci
Guideline values
For marine recreational waters used for primary contact activities, the guideline values are as
follows:
Geometric mean concentration (minimum of five samples):
≤ 35 enterococci/100 mL
Single-sample maximum concentration:
≤ 70 enterococci/100 mL
Calculation of the geometric mean concentration should be based on a minimum of five samples,
collected at appropriate times and sites to provide representative information on the water quality
likely to be encountered by users. Further action should be initiated if either of these guideline
values is exceeded. Minimum action should consist of immediate resampling of the site (or
sites). In addition, a swimming advisory may be issued should the responsible authority identify
that the area is not suitable for recreational water use.
It is further advised that recreational water areas routinely used for primary contact recreation be
monitored at a minimum of once per week, with increased monitoring recommended for those
beaches that are highly frequented or are known to experience high user densities. Similarly,
under certain scenarios, a reduction in the recommended sampling frequency may be justified.
Further guidance on sampling frequency recommendations and the posting of recreational waters
can be found in Part I (Management of Recreational Waters).
E. coli (Section 4.1.1) is also recognized as a useful predictor of the risk of gastrointestinal
illness in marine recreational waters (Wade et al., 2003). If it can be shown that E. coli can
adequately demonstrate the presence of faecal contamination in marine waters, then the E. coli
maximum limit for fresh waters may be adopted. If there is any doubt, samples should be
examined for both sets of indicators for extended periods to determine whether a positive
relationship exists.
Guideline rationale
The guideline values have been developed based on epidemiological evidence relating
enterococci concentrations in marine recreational waters to the incidence of swimmingassociated gastrointestinal illness observed among swimmers. The existing epidemiological data
are not sufficient to permit the estimation of the level of risk for individual exposures. Based on
the U.S. EPA’s regression analysis of epidemiological data (Cabelli, 1983), Health Canada has
estimated that using the guideline values for the recommended indicators of faecal contamination
for fresh and marine waters will correspond to a seasonal gastrointestinal illness rate of
approximately 1–2% (10–20 illnesses per 1000 swimmers). In Canada, owing to geography and
climate, a significantly smaller percentage of the population engages in marine recreational water
activities compared with those engaging in freshwater recreation.
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Guidelines for Canadian Recreational Water Quality: Third Edition
In determining the value for the maximum faecal indicator concentration permitted in a single
sample, the U.S. EPA equations pertaining to the calculation of a single-sample limit were
reviewed (U.S. EPA, 1986). The data regarding the maximum allowable indicator density at
designated beach areas are consistent with applying a factor of 2 times the recommended
geometric mean value results. Subsequently, a single-sample maximum concentration of 70
enterococci/100 mL is reaffirmed.
These values represent risk management decisions that have been based on a thorough
assessment of the potential risks for the recreational water user. In considering both the potential
health risks and the benefits of recreational water use in terms of physical activity and
enjoyment, it was concluded that this is a tolerable and reasonable estimate of the risk of illness
likely to be experienced by users engaged in a voluntary activity.
Evaluation of the epidemiological information published since the Guidelines were last issued
concluded that the current body of evidence supports the existing recommendations regarding the
use of enterococci as the indicator of faecal contamination in marine recreational waters. There
has not been substantial evidence to suggest that revision of the existing guideline values is
necessary at this time.
Description
Enterococci are members of the genus Enterococcus. The genus was created to include the more
faecal-specific species of genus Streptococcus, formerly considered as group D streptococci. In
practice, the terms enterococci, faecal streptococci, Enterococcus and intestinal enterococci have
been used interchangeably (Bartram and Rees, 2000). Enterococci are characterized by their
ability to fulfil the following criteria: growth at temperatures of 10°C and 45°C, resistance to
60°C for 30 minutes, growth in the presence of 6.5% sodium chloride and at pH 9.6, and the
ability to reduce 0.1% methylene blue (Bartram and Rees, 2000; APHA et al., 2005). Species of
the genus include E. faecalis, E. faecium, E. durans, E. hirae, E. gallinarum and E. avium.
E. faecalis and E. faecium occur in significant quantities in both human and animal faeces and,
along with E. durans, have been reported to be the species most frequently encountered in
polluted aquatic environments (Bartram and Rees, 2000). E. gallinarum and E. avium occur at
high concentrations in animal faeces, but are not exclusively associated with animal sources.
Enterococci closely satisfy many characteristics of a suitable indicator of faecal contamination in
recreational waters. Many species within the group enterococci are found in high numbers in
human and animal faeces. They are not commonly found in unpolluted waters and are generally
regarded to be incapable of growth in recreational waters (Ashbolt et al., 2001). Compared with
other indicator organisms (e.g., E. coli, thermotolerant coliforms), enterococci have
demonstrated greater resistance to certain environmental stresses in recreational waters, such as
conditions of sunlight and salinity. Enterococci have also demonstrated greater resistance to
wastewater treatment practices, including chlorination. A strong correlation has also been
demonstrated between the concentration of enterococci in marine waters and the risk of
gastrointestinal illness among swimmers (Cabelli, 1983; Kay et al., 1994).
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Guidelines for Canadian Recreational Water Quality: Third Edition
In the past, a ratio of thermotolerant coliforms to faecal streptococci concentrations was used in
attempts to indicate the origin of bacterial contamination (Geldreich, 1976; Clausen et al., 1977).
A thermotolerant coliform/faecal streptococcus ratio of 4 or higher was said to indicate a human
source of contamination, whereas a lower ratio would represent an animal source. However,
because of the noted differences in the survival times between these two groups in the
environment and the variability of the different methods used for their enumeration, the use of
the thermotolerant coliform/faecal streptococcus ratio is now considered inaccurate (Ashbolt et
al., 2001; APHA et al., 2005). As a result, the use of this ratio is not recommended. Further
information on faecal pollution source tracking can be found in Section 10.0 (Faecal pollution
source tracking).
Occurrence in the aquatic environment
Enterococci can be routinely isolated from marine and fresh recreational waters known to be
impacted by human and animal faecal pollution sources. These organisms are present in high
concentrations in human and animal faeces, with concentrations reported on the order of 106–
107/g (Sinton, 1993; Edberg et al., 2000). Overall, it is thought that enterococci are present at
concentrations approximately 1- to 3-fold lower than those of E. coli in faeces and municipal
wastes (Sinton, 1993; Edberg et al., 2000). Human faecal flora studies reported by Leclerc et al.
(2001) demonstrated that Enterococcus species could be detected in the faeces of 100% of the
subjects tested.
Several publications have reported that prolonged survival of enterococci is possible in marine
and freshwater sediments (Davies et al., 1995; Desmarais et al., 2002; Ferguson et al., 2005).
These are thought to provide more favourable conditions of temperature and nutrients than the
adjacent recreational waters. Others have reported on the ability of enterococci to survive in
organic-rich environments not known to be associated with faecal contamination, such as on
mats of the green algae species Cladophora (Whitman et al., 2003).
In Canada, there have been few published investigations on the distribution of enterococci in the
marine environment. Gibson and Smith (1988) conducted a study to investigate the distribution
of enterococci at 26 marine beaches in the Vancouver region. The findings of this study
demonstrated that 1.6% of the results would have exceeded the enterococci geometric mean
concentration guideline value of 35/100 mL. In 1988, the New Brunswick Department of Health
and Community Services (1989) monitored eight marine beaches along the Northumberland
Strait in New Brunswick. The overall enterococci levels were low, showing a geometric mean
concentration of 3.5/100 mL. The results of the study indicated that enterococci were absent in
60% of the samples.
Association with pathogens
Enterococci are considered a good indicator for enteric bacterial pathogens. In a survey of
surface waters collected from various watersheds in southern California, enterococci were shown
to demonstrate good predictive ability with PCR detection of STEC (Yanko et al., 2004). It was
reported that above an enterococci concentration of 100 most probable number (MPN)/100 mL,
the probability of detection of STEC was approximately 60–70%.
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Guidelines for Canadian Recreational Water Quality: Third Edition
Enterococci are somewhat less effective as an indicator of the presence of enteric pathogenic
viruses and protozoa. A number of researchers have reported on the lack of a relationship
between enterococci concentrations and the presence of human viruses in surface waters
(Griffin et al., 1999; Schvoerer et al., 2000, 2001; Jiang et al., 2001, Jiang and Chu, 2004)
Enterococci are considered the best available indicator of water quality for marine recreational
waters (Pruss, 1998; WHO, 1999; Wade et al., 2003). Detection indicates faecal contamination
of water and thus the possible presence of faecal pathogenic bacteria, viruses and protozoa.
Human enteric viral and protozoan pathogens of faecal origin can survive for prolonged periods
in marine waters. Although the presence of high enterococci counts may indicate the possible
presence of viral and protozoan pathogens, the opposite—that the absence of enterococci
indicates that these pathogens are also absent—cannot be assured.
The combination of routine monitoring for enterococci alongside actions, procedures and tools to
collectively reduce the risk of swimmer exposure to faecal contamination in the recreational
water environment represents the most effective approach to protecting the health of recreational
water users.
Guidelines used by other countries/organizations
The guideline formats and values established by other government and multinational
organizations worldwide for faecal indicator organisms in marine waters were reviewed in
developing the revised edition of this document.
Table 3. Guideline values for faecal indicator concentrations in marine recreational waters established by other
countries or organizations.
Country/
Marine water
Format and guideline values
Reference
organization
indicator
U.S. EPAa
Enterococci
Geometric mean concentration:
U.S EPA, 2002
35/100 mL
Single sample maximum concentration:b
104/100 mL
WHO
Intestinal
95th percentile/100 mL:
WHO, 2003a
enterococci
A: ≤40
B: 41-200
C: 201-500
D: >500
Australia
Intestinal
95th percentile/100 mL:
NHMRC, 2008
enterococci
A: ≤40
B: 41-200
C: 201-500
D: >500
European Union
Intestinal
95th percentile/100 mL:
EU, 2006
enterococci
Excellent: 100 /100 mL
Good: 200/100 mL
E. coli
90th percentile/100 mL:
Sufficient: 185/100 mL
95th percentile/100 mL:
Excellent: 250 /100 mL
Good 500/100 mL
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Guidelines for Canadian Recreational Water Quality: Third Edition
90th percentile/100 mL:
Sufficient: 500/100 mL
a – Note that new criteria is currently under development and may be available in 2012.
b – Designated beach area (75% confidence level)
Related epidemiological studies
The U.S. EPA’s original epidemiological studies in marine recreational waters (Cabelli, 1983)
measured the concentrations of faecal indicator organisms - total coliforms, thermotolerant
(faecal) coliforms, E. coli, enterococci - in swimming waters and compared them with the rates
of swimming-associated illness reported on the same days on which the samples were collected.
Statistically significant rates of gastrointestinal illness were observed among swimmers in waters
considered to be more faecally contaminated. For symptoms unrelated to gastrointestinal illness,
no statistically significant differences were observed. For the data analysis, the mean seasonal
faecal indicator concentration per 100 mL was plotted against the seasonal swimming associated
rate of gastrointestinal symptoms per 1000 persons. Correlation and regression analysis were
then used to determine the correlation coefficients and the slope of the linear regression equation
for each indicator. The best correlation coefficient (r) was obtained with enterococci (r = 0.75).
The following regression equation was produced for the enterococci data :
seasonal risk of gastrointestinal
illness per 1000 persons
= 0.20 + 12.17(log enterococci concentration per 100 mL).
Several marine water epidemiological studies have been conducted since the Guidelines for
Canadian Recreational Water Quality were last developed (Cheung et al., 1990; Alexander et
al., 1992; von Schirnding et al., 1992; Corbett et al., 1993; Harrington et al., 1993; Kay et al.,
1994; Kueh et al., 1995; Marino et al., 1995, Fleisher et al., 1996; van Dijk et al., 1996; McBride
et al., 1998; Haile et al., 1999; Prieto et al., 2001). All of the studies have confirmed the
existence of a relationship between exposure to marine recreational waters of varying quality and
symptoms of water-related illness among swimmers. The most significant findings came from
the results of the randomized controlled program of epidemiological studies conducted at coastal
beaches in the United Kingdom (Kay et al., 1994; Fleisher et al., 1996). These studies were
designed to address some of the perceived shortcomings of the traditional beach survey design
used in many of the earlier studies. The key features of the randomized controlled design were
efforts to ensure a more random distribution of subjects in the swimming and non-swimming
groups and tighter monitoring of the water quality experienced by the individual swimmers. Of
the faecal indicators monitored, only faecal streptococci levels measured at chest depth showed a
significant relationship with the incidence of both gastrointestinal illness and respiratory illness
among swimmers. The authors further reported the existence of possible thresholds for an
increased risk of gastroenteritis at a concentration of 32 faecal streptococci/100 mL and an
increased risk of respiratory illness at a concentration of 60 faecal streptococci/100 mL. In other
studies, McBride et al. (1998) reported an increasing risk of respiratory illness with increasing
enterococci levels among swimmers at New Zealand beaches. Cheung et al. (1990) observed a
moderate correlation (r = 0.63) between enterococci levels and rates of highly credible
gastrointestinal illness (HCGI) and skin symptoms combined at coastal beaches in Hong Kong,
although a stronger correlation was observed with E. coli (r = 0.73).
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Guidelines for Canadian Recreational Water Quality: Third Edition
A few epidemiological studies have been conducted that have investigated the health effects of
recreational activities other than swimming, such as surfing (Gammie and Wyn-Jones, 1997;
Dwight et al., 2004). The data linking water quality and illness through these activities have been
less strong. Still, conclusions that have been reported are that gastrointestinal illness similarly
constitutes the most frequently reported health outcome during these types of activities and that
factors related to the risk of illness include the quality of the water and the frequencies of
immersion and water ingestion.
Several reviews of the epidemiological findings have also been published. WHO (Pruss, 1998)
published a comprehensive review of all of the recreational water epidemiological studies
conducted over the period from 1953 to 1996. From the review, it was concluded that
gastrointestinal symptoms were the most frequently reported outcome for which water quality
dose–response relationships were reported, and that the indicator organisms that best correlated
with the health outcomes were enterococci for marine waters and E. coli and enterococci for
fresh waters. The U.S. EPA has also published two reviews of the existing epidemiological
literature for recreational waters. The first, published in its Implementation Guidance for
Ambient Water Quality Criteria for Bacteria (U.S. EPA, 2002), was a minireview of the
epidemiological studies conducted since the previous guidance had been issued, in 1986. The
U.S. EPA concluded that the epidemiological methods used to derive its 1986 Water Quality
Criteria remained scientifically valid and that no new scientific principles had been established
that would require a revision of the current guidelines. More recently, Wade et al. (2003)
conducted a meta-analysis of all of the existing epidemiological data published since 1950
linking microbiological indicators of recreational water quality to gastrointestinal illness in
swimmers. The authors concluded that in the marine water studies, enterococci and, to a lesser
extent, E. coli were the most reliable predictors of gastrointestinal illness. Moreover, the authors
observed that the reported risks of gastrointestinal illness at enterococci concentrations below the
current U.S. EPA standards were not statistically significant, whereas values above the standards
were elevated and statistically significant.
The U.S. EPA and CDC are currently conducting epidemiological studies at freshwater and
marine beaches under the NEEAR Water Study. The studies are intended to support the
development of new recreational water quality guidelines (U.S. EPA, 2002; Dufour et al., 2003),
as well as to investigate new water quality indicators and rapid methods for water quality
monitoring. Data collection for these studies was completed in 2010.
Summary
Based on all of the existing evidence, the enterococci group remains the most suitable indicator
of faecal contamination in marine recreational waters. In summary:
1. The guideline values have been developed based on the analysis of epidemiological evidence
relating enterococci concentrations in marine recreational waters to the incidence of
swimming-associated gastrointestinal illness observed among swimmers. The values
represent risk management decisions based on the assessment of possible health risks for the
recreational water user and the recognition of the tremendous benefits that recreational water
activities provide in terms of health and enjoyment.
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Guidelines for Canadian Recreational Water Quality: Third Edition
2. Enterococci most closely fit the requirements of an ideal indicator of faecal contamination
for marine recreational waters. Detection suggests faecal contamination of water and thus the
possible presence of faecal pathogenic bacteria, viruses and protozoa.
3. There are limitations associated with the use of indicators in assessing the quality of
recreational waters. The occurrence of faecal pathogens in recreational waters is dependent
on many factors and can be variable and sporadic. The absence of enterococci should not be
interpreted to mean that enteric pathogenic microorganisms are also absent.
4. Combining routine enterococci monitoring alongside actions, procedures and tools to
collectively reduce the risk of swimmer exposure to faecal contamination in the recreational
water environment represents the most effective approach to protecting the health of
recreational water users.
4.2
Advice regarding water intended for secondary contact recreational activities
The Guidelines for Canadian Recreational Water Quality are intended to be protective for those
activities that involve intentional or incidental immersion in natural waters. Due to increased
interest from jurisdictions in distinguishing between primary contact activities and secondary
contact activities, this current edition of the Guidelines takes an initial step at providing advice
for secondary contact activities separately from that of primary contact with respect to faecal
indicator concentration.
Recreational activities that have been traditionally considered as secondary contact activities
(e.g. canoeing, fishing) involve exposures much different from those associated with primary
contact uses. Ingestion of water and, subsequently, the risk of gastrointestinal illness are
presumed to be lower during secondary contact recreation. Still, it is expected that there is some
degree of risk of acquiring illness through these activities. Inadvertent immersion can result in
whole body contact, and splashing can lead to a variety of water exposure scenarios. Illnesses
affecting the skin and perhaps the mucous membranes of the eyes and ears may be of relatively
greater importance for secondary contact uses (U.S. EPA, 2002). Inhalation may also be an
important route of exposure during secondary contact activities in areas where sprays or aerosols
are generated.
Limited research has been conducted on the potential risks of acquiring illness during secondary
contact activities in recreational waters. In one study which investigated the relationship between
water quality and illness acquired during canoeing or rowing, Fewtrell et al. (1994) noted no
significant differences between the exposed group and the unexposed group. The bulk of the
epidemiological research on recreational water uses and the risk of acquiring illness have been
generated for primary contact activities. As a result, insufficient data are available to derive
precise health-based faecal indicator limit values intended to protect users engaged in secondary
contact recreational activities from exposure to faecal contamination. However, it is recognized
that because a lower degree of water exposure can be expected at most times during the majority
of secondary contact recreational activities, there may be some water areas where a secondary
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Guidelines for Canadian Recreational Water Quality: Third Edition
contact use designation with separate water quality values is desired and considered reasonable
and acceptable to local and regional authorities.
When contemplating the establishment of separate faecal indicator values for water areas used
entirely for secondary contact recreational uses, a clear understanding of the types of activities
that would be considered to fit under this description is required. WHO, in its Guidelines for Safe
Recreational Water Environments: Volume 1—Coastal and Fresh Waters (WHO, 2003a), has
proposed a scheme for the classification of recreational water activities according to their degree
of water exposure. The following descriptions (adapted from WHO, 2003a), may be used as an
initial guide when determining whether a specific recreational activity would be considered as
either primary or secondary contact:
•
•
Primary contact: Recreational activity in which the whole body or the face and trunk are
frequently immersed or the face is frequently wetted by spray, and where it is likely that
some water will be swallowed. Inadvertent immersion, through being swept into the water by
a wave or slipping, would also result in whole body contact. Examples include swimming,
surfing, waterskiing, whitewater canoeing/rafting/kayaking, windsurfing or subsurface
diving.
Secondary contact: Recreational activity in which only the limbs are regularly wetted and in
which greater contact (including swallowing water) is unusual. Examples include rowing,
sailing, canoe touring, or fishing.
Even if these classification criteria are used, it remains a significant challenge to discern which
activities constitute primary contact and which constitute secondary contact. The classification of
certain recreational water activities will be clear, whereas that of others may be less obvious and
more open to interpretation. Activities considered as potential candidates under a secondary
contact use designation should be evaluated on a case-by-case basis.
Other factors to consider before assigning a secondary contact use designation to a recreational
water area include:
• The water area should first be subject to an assessment of existing uses, water quality and
the potential for improvement as well as any other relevant factors, such as health or
environmental considerations.
• The secondary contact designation should not be applied where an assessment has shown
primary contact recreation to be a significant use.
• Where the water area has a shared use (e.g. swimming and canoeing), it is the primary
contact values that should apply.
• When an area is posted as suitable only for secondary contact recreational uses,
communication material should clearly convey that accidental immersion (through falls,
canoe spills, etc.) can lead to whole body exposure; under these circumstance, water
ingestion may result in illness.
• Users should be reminded to take the precautions necessary to ensure that these types of
exposures are avoided as much as possible; the skill of the person performing the activity
may strongly influence the degree of water exposure.
• Further guidance on the posting of information at recreational water areas can be found in
Part I (Management of Recreational Waters).
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Guidelines for Canadian Recreational Water Quality: Third Edition
Responsible authorities have a duty to take precautions, protect the health and safety of all
recreational water users and maintain the best water quality possible. The existence of the
secondary contact values should not be used as a mechanism for downgrading the status of an
area in response to poor water quality issues. This is particularly important where an assessment
has shown that the primary contact guideline values could be achieved.
Based on an assessment of the available information, where a water area is intended to be used
specifically for secondary contact recreation (i.e. where primary contact is not an existing use),
the application of a factor of 5 to the geometric mean faecal indicator concentration used to
protect primary contact recreation users may be used as an approach to establish faecal indicator
limits. The corresponding values for E. coli and enterococci concentrations would therefore be as
follows:
Fresh waters – E. coli: (5 × 200/100 mL) = 1000 E. coli/100 mL
Marine waters – Enterococci: (5 × 35/100 mL) = 175 enterococci/100 mL
These values represent a risk management decision based on the assessment of the expected
exposure scenarios and potential health risks for the recreational water user. They are intended
to allow specified water areas to have a secondary contact use designation where this has been
considered appropriate by the responsible local or regional authorities, while still providing some
level of protection for secondary contact recreational users until epidemiologically based
guideline values can be derived. In considering both the potential health risks and the benefits of
recreational water use, it was concluded that this is a tolerable and reasonable approach to protect
users engaged in a voluntary activity. These values are also consistent with advice provided by
other jurisdictions (Saskatchewan Environment, 1997; Alberta Environment, 1999; U.S. EPA,
2002; British Columbia Ministry of Health Services, 2007; MDDEP, 2007). The values will be
periodically reviewed or adjusted, however, as new or more significant data become available.
Insufficient information is availableto develop separate values with respect to water areas used
solely for secondary-contact for other parameters in the Guidelines for Canadian Recreational
Water Quality. Area operators, service providers and responsible authorities should remain
aware that these parameters may also affect water areas intended only for secondary contact
uses. Where separate guidance for secondary contact does not exist, it is advised that the values
and associated guidance provided in the Guidelines apply to all recreational waters, regardless of
the types of activities practised.
4.3
Other organisms as potential indicators
Many different waterborne pathogens can be encountered in Canadian recreational waters. As
discussed in the previous section, recreational water quality is most frequently determined by
the presence of faecal indicator bacteria which, by extension, suggests the possible existence of
faecally transmitted waterborne pathogens. Currently, E. coli (fresh waters) and enterococci
(marine waters) remain the best available indicators of recreational water quality, as they, above
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Guidelines for Canadian Recreational Water Quality: Third Edition
all other organisms, have been shown to be the most successful in meeting the desired indicator
criteria.
Nevertheless, the current indicators do not unequivocally meet all of the requirements of an ideal
indicator, and the limitations of these two organisms as pathogen indicators are well known. It is
understood that no single organism would be able to fill all of the roles of what might be
considered a perfect indicator of recreational water quality—one that models all of the known
pathogens, provides information on the degree and source of faecal contamination and
communicates the potential risk of illness for recreational water users. It has been proposed that
this task would require multiple indicators, each with unique characteristics that would enable
them to satisfy specific roles (Ashbolt et al., 2001).
The term indicator can be further specified to reflect these different functions. Indicators may
be considered as faecal indicators (indicative of the presence of faecal contamination, but not
necessarily of specific pathogens) or pathogen indicators (indicative of the presence and
behaviour of specific pathogens). Moreover, faecal indicators may be further categorized into
what can be considered primary indicators—those that provide information on the magnitude or
extent of faecal contamination— and secondary indicators—those that provide information on
the source of faecal contamination.
The objective of this section is to provide a summary of what is known regarding other
microorganisms that have been widely discussed among recognized experts, academics and
policymakers as potential indicators for recreational water. The organisms covered are
Bacteroides spp., Clostridium perfringens, F+ RNA coliphages and bacteriophages infecting
Bacteroides fragilis. A summary of the characteristics of the recommended indicators and the
other potential indicator organisms discussed is presented in Table 4.
Potential indicators
Bacteroides spp.
Bacteroides spp. are Gram-negative, rod-shaped, obligate anaerobic bacteria. The genus is
considered to represent the most abundant bacterial genus in human faeces (Fiksdal et al., 1985).
The four dominant species—B. fragilis, B. vulgatus, B. distasonis and B. thetaiotaomicron—can
reach concentrations on the order of 1010 cells/g faeces (Kator and Rhodes, 1994), outnumbering
E. coli concentrations by as much as 100- to 1000-fold (Slanetz and Bartley, 1957; Holdeman et
al., 1976; Fiksdal et al., 1985). Bacteroides species can occur in much lower densities (105- fold
to 1010-fold lower) in animals (Allsop and Stickler, 1985; Kator and Rhodes, 1994), although
higher densities have been observed in some species, such as domestic pets and gulls (107–108
cfu/g) (Allsop and Stickler, 1985).
Because of their high numbers in human faeces, Bacteroides species have long been considered
a candidate indicator of faecal pollution; however, the difficulties associated with culturing
anaerobic bacteria discouraged their use in investigations (Kreader, 1995; Bernhard and Field,
2000a). Recent advances in molecular biology have overcome this problem. Researchers have
developed PCR assays for the detection of both generic and species-specific (human, bovine)
Bacteroides genetic markers in faeces. In this case, the presence of the genetic markers is
considered to infer the presence of Bacteroides cells.
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Guidelines for Canadian Recreational Water Quality: Third Edition
Bacteroides PCR methods have been successfully used for the detection of faecal pollution in
contaminated water samples (Kreader, 1998; Bernhard and Field, 2000b; Field et al., 2003). As
well, quantitative PCR (QPCR) methods have been developed that have enabled the near realtime enumeration of Bacteroides spp. in recreational waters (Fung, 2004; Seurinck et al., 2005;
Wade et al., 2006). The U.S. EPA included QPCR monitoring for Bacteroides as part of its
NEEAR Waters Study (Wade et al., 2006).
Detection of Bacteroides genetic markers in recreational waters is a relatively new area of
research. Few studies have been conducted to date that provide an analysis of Bacteroides
markers relative to faecal indicator organisms, faecal pathogens or rates of illness among
swimmers. Wade et al. (2006) reported a positive but weak association between Bacteroides and
gastrointestinal illness among swimmers at one of two freshwater beaches investigated during
the NEEAR Waters Study. A problem with the sensitivity of the QPCR method was cited by the
authors (Wade et al., 2006).
Noted strengths of Bacteroides as a possible primary indicator of faecal contamination include
their high occurrence in human faeces and sewage, the inability of the bacterium to grow in the
environment and the long environmental persistence of the DNA markers. Limitations include
the lower concentrations observed in non-human faecal sources, current data gaps regarding its
capabilities as an indicator (primary indicator, pathogen indicator and indicator of swimmingassociated illness) and challenges relating to the analytical methods (expensive, technically
demanding, issues of sensitivity).
Information generated to date suggests that Bacteroides markers may have a stronger role as a
secondary indicator of faecal contamination, providing information on the potential sources of
faecal material.
Clostridium perfringens
C. perfringens is a Gram-positive, rod-shaped, spore-forming anaerobic bacterium that is
consistently found in both human and non-human faeces (Bisson and Cabelli, 1980). Clostridium
species have the ability to enter into a protective spore form that is able to resist environmental
stresses and that can allow the organism to persist in the environment for long periods of time.
C. perfringens has long been considered a valuable indicator of the sanitary quality of water,
dating back to the late 19th century (Ashbolt et al., 2001). Much of the interest in the use of
C. perfringens as an indicator of recreational water quality has stemmed from research conducted
in the state of Hawaii. Researchers there had observed that many stream and soil samples not
thought to be affected by a known source of faecal pollution contained faecal coliforms and
E. coli in excess of the current water quality standards (Fujioka and Shizumura, 1985). It was
also observed that C. perfringens concentrations in streams receiving wastewater discharges
were consistently higher than in streams that were not affected (Fujioka and Shizumura, 1985).
The researchers subsequently asserted that C. perfringens was a more reliable indicator of faecal
contamination in Hawaiian waters. It has since been proposed that this situation may be
encountered in other tropical locations in the United States (U.S. EPA, 2001b). At present,
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Guidelines for Canadian Recreational Water Quality: Third Edition
Hawaii is the only known jurisdiction that includes monitoring for C. perfringens as a
recreational water quality indicator (Anon., 1996).
The concentration of C. perfringens in human and animal faeces is considerably less than that of
E. coli or enterococci (Wright, 1982). Published data suggest that C. perfringens may be detected
in only a small to moderate percentage of human faecal samples (13–35%), with an average
faecal concentration of approximately 103 cells/g (Ashbolt et al., 2001). Higher concentrations of
C. perfringens have been reported in sewage (Fujioka and Shizumura, 1985). C. perfringens has
been detected in the faeces of a wide range of animals, including birds, mammals, reptiles and
amphibians (Conboy and Goss, 2003). Significant numbers of the organism have been
encountered in the faeces of a few animal species, including dogs (108 cells/g), cats (107 cells/g)
and sheep (105 cells/g) (Ashbolt et al., 2001). C. perfringens is not exclusively associated with
faecal wastes and is a common inhabitant of soils (Toranzos, 1991).
Water quality investigations conducted in coastal waters in Florida have indicated that
concentrations of C. perfringens do not correlate well with levels of faecal indicator bacteria
(Griffin et al., 1999; Lipp et al., 2001) or with the presence of enteric viruses (Griffin et al.,
1999). Furthermore, the levels of C. perfringens were generally lower than those of enterococci
or the faecal coliform group in the water column, but were substantially greater than the levels of
either indicator in the underlying sediment (Lipp et al., 2001). In a study of indicator and
pathogen presence in selected lakes and rivers in southwestern Finland, Hörman et al. (2004)
reported a positive correlation between the presence of C. perfringens and the detection of one or
more of the pathogens being tested for (Cryptosporidium, Giardia, Campylobacter, noroviruses).
However, the absence of C. perfringens demonstrated only a weak predictive value for a
negative pathogen sample.
Several epidemiological studies have included C. perfringens as an indicator when investigating
the relationships between water quality and illness in swimmers (Cabelli, 1983; Harrington et al.,
1993; Kueh et al., 1995; Lee et al., 1997; Wiedenmann et al., 2006). Cabelli (1983) observed a
weak correlation between C. perfringens densities and acute gastrointestinal illness in swimmers
during the U.S. EPA’s original epidemiological studies at marine beaches in the 1970s. Kueh et
al. (1995) noted a positive, but not strong, correlation with the incidence of gastroenteritis among
swimmers and C. perfringens in two marine beaches in Hong Kong. Wiedenmann et al. (2006)
did report a relationship between C. perfringens and the incidence of gastroenteritis among
swimmers during a randomized controlled epidemiological study at German freshwater beaches.
A NOAEL of 13 C. perfringens/100 mL was reported for several definitions of gastrointestinal
illness.
Strengths of C. perfringens include its inability to grow in the environment and its capability for
surviving longer than faecal waterborne pathogens. Advances in culture methods (Adcock and
Saint, 2001) have helped to improve the ease with which C. perfringens can be detected—
previously an impediment surrounding the use of this organism as a water quality indicator.
Limitations include not being faecal specific, having lower faecal numbers relative to other
indicator bacteria; detection being strongly dependent on the source of contamination; presence
not necessarily indicative of recent contamination (owing to the long environmental persistence
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Guidelines for Canadian Recreational Water Quality: Third Edition
of the spores); and a lack of epidemiological evidence linking C. perfringens concentrations to
the potential for acquiring swimming-associated illness.
C. perfringens may be better suited as an indicator of the effectiveness of drinking water
treatment processes (Bisson and Cabelli, 1980; Payment and Franco, 1993) or as an indicator of
intermittent or cumulative sewage inputs (Sorensen et al., 1989; Hill et al., 1993; Lisle et al.,
2004). At present, C. perfringens appears to more closely satisfy the role of a pathogen indicator
or perhaps a secondary indicator of faecal contamination.
F+ RNA coliphages
Coliphages are bacteriophages (viruses infecting only bacteria) that specifically infect E. coli
cells. The reasons cited for investigating coliphages as potential indicators of faecal
contamination are that coliphages are thought to more closely resemble enteric viruses in terms
of their physical characteristics, environmental persistence and resistance to disinfection,
compared with the traditional bacterial indicators of faecal contamination. They are also less
costly and easier to enumerate than human viruses. Also, because they theoretically infect only
E. coli cells, it is thought that their detection should be sufficiently indicative of the presence of
faecal contamination.
There are two main types of coliphages: somatic coliphages and male-specific (F+) coliphages.
Somatic coliphages infect E. coli cells by attaching to the lipopolysaccharide component of the
cells’ outer membranes. They have been investigated as potential indicators of swimming water
quality (Contreras-Coll et al., 2002; Vantarakis et al., 2005; Wiedenmann et al., 2006); however,
they are thought to represent a less specific group than the F+ coliphages, and knowledge of their
sources and behaviour is currently lacking. By comparison, F+ coliphages have been more
extensively studied (Duran et al., 2002).
F+ coliphages demonstrate a greater specificity than somatic coliphages, infecting E. coli cells
possessing F-pili—tube-like structures coded for by an F-plasmid and that allow connections to
form between cells for the transfer of genetic material (Singleton and Sainsbury, 1997; Scott et
al., 2002). It is these F-pili that serve as the site of phage attachment.
F+ coliphages include F+ RNA phages and F+ DNA phages. F+ RNA phages more closely
resemble the human viruses of waterborne significance and have thus been preferentially
explored (Sobsey, 2002). Using immunological or genetic methods, the group can be further
categorized into four distinct serogroups or genogroups, and subsequent source tracking
investigations have identified that the presence of a particular subgroup has some merit in
distinguishing the source of faecal contamination (Havelaar et al., 1990; Brion et al., 2002;
Schaper et al., 2002b; Cole et al., 2003). In general, Groups II and III have been shown to be
highly associated with human faecal contamination (e.g., domestic or municipal sewage), Group
IV has been shown to be predominantly linked to animal faecal contamination and animal waste
materials, Group I has been isolated from both human and animal faecal material and wastes
(Scott et al., 2002; Sobsey, 2002).
F+ RNA phages are not always present in human faeces and, when detected, are often present in
low numbers (Havelaar and Pot-Hogeboom, 1988; Havelaar et al., 1990; Luther and Fujioka,
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Guidelines for Canadian Recreational Water Quality: Third Edition
2004). Researchers have similarly reported low isolation frequencies among septage samples or
waters known to be affected by septic system wastes (Calci et al., 1998; Griffin et al., 1999). A
low isolation frequency among animal faecal samples has also been reported (Calci et al., 1998;
Luther and Fujioka, 2004). Significantly higher numbers of F+ RNA coliphages have been
detected in sewage and wastewater (Contreras-Coll et al., 2002; Lucena et al., 2003).
Sinton et al. (1999) reported that the degree of survival for various indicator organisms in
sewage-polluted seawater during sunlight inactivation experiments (simulated summer
conditions) was (from greatest to least): somatic coliphages > F+ RNA phages > enterococci
> E. coli. The F+ RNA coliphage groups have been shown to vary markedly in their ability to
persist in the environment (Brion et al., 2002; Schaper et al., 2002a; Sobsey, 2002; Long and
Sobsey, 2004). In general, it has been observed that Group I is capable of the longest
environmental persistence, followed by Groups II and III, with Group IV demonstrating the
shortest period of survival (Brion et al., 2002; Schaper et al., 2002a; Long and Sobsey, 2004).
Evidence as to whether F+ RNA coliphages are a reliable indicator of faecal pollution in natural
waters has been conflicting. Havelaar et al. (1993) and Ballester et al. (2005) reported that
F+ RNA coliphage concentrations were more strongly correlated with concentrations of
infectious enteroviruses and enteric viruses than either faecal coliforms or enterococci in
environmental waters. The authors did report, however, that in a few instances, viruses were
isolated in the absence of the coliphages, and that the reverse was also true (Havelaar et al.,
1993). Griffin et al. (1999) demonstrated a lack of predictive ability when using coliphages
(somatic and F+ RNA) as indicators of the presence of enteric viruses in canal waters in the
Florida Keys. Jiang and Chu (2004) reported no apparent relationship between the detection
of adenovirus, enterovirus and hepatitis A virus (HAV) genomes and F-specific coliphage
concentrations in a study of human viral contamination in river and coastal waters in southern
California.
Grabow (2001) suggested that a direct correlation between the number of coliphages and enteric
viruses in water environments cannot be expected, as coliphages are excreted at all times by a
certain percentage of the human population, whereas enteric viruses are largely excreted during
infection, which can be intermittent and seasonal.
There has been some information collected regarding the relationship between coliphage
concentrations and the rate of swimming-associated illness. Lee et al. (1997) reported a
significant association between the concentration of F+ RNA coliphages and the reporting of
gastrointestinal symptoms among canoeists and rafters at an artificial freshwater canoe course in
the United Kingdom. Other researchers have included coliphages among the panel of prospective
indicators monitored during epidemiological investigations (von Schirnding et al., 1992; Marino
et al., 1995; McBride et al., 1998; van Asperen et al., 1998); however, no significant correlations
were identified.
Strengths of F+ RNA phages as a possible faecal indicator include the noted similarities to human
enteric viruses, strong evidence of being exclusively associated with human and animal faecal
material, an apparent inability to replicate in the environment and potential applications in faecal
source identification. Limitations include presence being dependent on the source of
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Guidelines for Canadian Recreational Water Quality: Third Edition
contamination, different survival rates of the individual phage groups and a lack of a
demonstrated relationship with either enteric virus presence or rates of swimming-associated
illness.
F+ RNA coliphages appear to be more useful as an indicator of sewage, rather than of faecal
contamination in general. At present, they appear to be better positioned as potential pathogen
indicators or secondary indicators, as opposed to primary indicators of faecal contamination
(Chapron et al., 2000).
Bacteriophages of Bacteroides fragilis
Bacteriophages of B. fragilis have been investigated as a possible indicator of faecal
contamination on the theory that the organism might combine some of the desirable properties
reported for both the coliphage group and Bacteroides spp.—the potential to be present in high
numbers in faecal material, while demonstrating survival traits more representative of enteric
viruses.
Although concentrations of B. fragilis in human faeces are high, B. fragilis phages have been
shown to be isolated somewhat infrequently in faeces, and in lower numbers (Tartera and Jofre,
1987; Grabow et al., 1995). Published accounts have placed the percentage of human faecal
samples from which B. fragilis phages have been isolated in the range from 10% to 28% (Tartera
and Jofre, 1987; Grabow et al., 1995; Puig et al., 1999; Gantzer et al., 2002). Phage isolation
from human and animal faeces has been shown to be largely dependent upon the B. fragilis host
strain used for recovery (Tartera and Jofre, 1987; Puig et al., 1999). Molecular methods for the
detection of B. fragilis bacteriophages are under investigation, and these are expected to
eliminate many of the difficulties associated with the current host recovery methods (Puig et al.,
2000).
B. fragilis phages have been shown to be routinely isolated from sewage; however, the
concentrations encountered (< 10–105 phages/100 mL) are frequently less than those reported for
somatic and F+ RNA coliphages (Puig et al., 1999; Contreras-Coll et al., 2002; Lucena et al.,
2003). Tartera and Jofre (1987) reported that B. fragilis phages could be detected in all water and
sediment samples from highly polluted rivers (101–105 plaque-forming units [pfu]/100 mL), but
not in samples collected from areas not known to receive sewage pollution. In a more recent
investigation of the occurrence and levels of phages of B. fragilis in swimming waters
throughout Europe, Contreras-Coll et al. (2002) documented a median phage concentration well
below 10 pfu/100 mL, with 95% of the samples below 102 pfu/100 mL. Lucena et al. (2003)
reported similar numbers during a survey of water samples collected from selected rivers in
Europe and South America.
No correlation has been shown between the concentration of faecal indicator bacteria and levels
of B. fragilis phages in recreational waters (Tartera et al., 1989; Contreras-Coll et al., 2002;
Lucena et al., 2003). Certain studies have suggested that B. fragilis phages are reliable indicators
of viral contamination in treated wastewaters (Gantzer et al., 1998) and shellfish (Hernroth et al.,
2002; Formiga-Cruz et al., 2003). However, there has been limited information published to date
in which the occurrence and concentrations of enteric viruses and phages of B. fragilis in
recreational waters have been directly compared. In a study of river water samples where
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domestic sewage was cited as its primary contamination source, Tartera et al. (1989) reported
that B. fragilis phages could be consistently isolated from samples in which enteroviruses were
detected.
Of the organisms suggested as possible alternative indicators of faecal pollution, B. fragilis
phages have been perhaps the least well investigated. Potentially useful properties include the
absence of significant non-faecal sources, the inability to replicate in the environment and
structural similarities to the enteric viruses. Limitations include low faecal concentrations,
variable isolation depending on the contamination source and difficulties associated with
recovery of the organisms.
Currently, it is speculated that, similar to the F+ coliphages, B. fragilis phages may be more
suitable as indicators of sewage contamination—and thus as possible secondary indicators of
faecal contamination.
Summary
1. The faecal contamination of recreational waters and the associated risk for water users are a
complex issue. E. coli and enterococci are considered the best available indicators of faecal
contamination in recreational waters; however, no single organism is capable of satisfying all
of the roles of an ideal indicator. Multiple indicators may be required for a more complete
understanding of this subject.
2. The organisms most widely discussed as other potential recreational water indicators include
Bacteroides spp., C. perfringens, F+ RNA coliphages and bacteriophages infecting B. fragilis
(see Table 4).
3. At present, none of the proposed indicators investigated meet a sufficient number of the
requirements necessary to be successful as a routine indicator of recreational water quality.
None of these organisms has demonstrated a consistent correlation with the presence of
waterborne pathogens in recreational waters, nor is there evidence of a strong
epidemiological link between the occurrence of these organisms and the incidence of
illness among recreational water users.
4. Nevertheless, these organisms do possess certain unique properties that might enable them
to fulfil other roles as recreational water indicators. Currently, these organisms appear to be
better suited as possible pathogen indicators or as faecal source indicators. Advances in
detection and enumeration methods may help improve our understanding of these organisms
and the roles they may play in recreational water monitoring programs in the future.
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1
Table 4. Characteristics of recommended and potential indicator organisms
Indicator organism
Criteria
E. coli
Enterococci
C. perfringens
Bacteroides spp.
(genetic markers)
F+ RNA coliphages
B. fragilis phages
Found within the
intestinal tract of
humans and
warm-blooded
animals.
Present in high
numbers in human
and animal faeces.
Present in high
numbers in human
and animal faeces.
Low numbers in
human and animal
faeces (high in certain
animal species).
Higher in sewage.
Very high numbers in
human faeces. Low to
high numbers in
animal faeces (species
dependent).
Low numbers and
variable isolation
among human and
animal faeces. Higher
in sewage.
Low numbers and
variable isolation
among human and
animal faeces. Higher
in sewage.
Present in faecescontaminated
waters when
enteric pathogens
present, and in
greater numbers.
Good indicator of all
sources of faecal
contamination.
Typically present in
faecal material in
higher concentrations
than pathogens.
Good indicator of all
sources of faecal
contamination.
Typically present in
faecal material in
higher concentrations
than pathogens.
Dependent upon
source. Recovery
difficult at low levels
of contamination.
Dependent upon
source. Insufficient
data on correlation
with pathogens.
Dependent upon
source. Recovery
difficult at low levels
of contamination.
Dependent upon
source. Recovery
difficult at low levels
of contamination.
Incapable of
growth in the
aquatic
environment.
Generally regarded as
true. Evidence to
suggest select strains
may be capable of
growth in soil
environment if proper
conditions are met.
Generally regarded as
true. Evidence to
suggest select strains
may be capable of
growth in soil
environment if proper
conditions are met.
Anaerobic bacteria.
Unable to replicate in
aquatic environment.
Anaerobic bacteria.
Unable to replicate in
aquatic environment.
Thought not to be
capable of replication
in aquatic
environment.
Possibility for
replication in sewage.
Host bacteria are
anaerobic. Thought
not to be capable of
replication in aquatic
environment.
Capable of
surviving longer
than pathogens.
Considered a good
indicator of survival
of pathogenic enteric
bacteria, but not
enteric viruses or
protozoa.
Considered a good
indicator of survival
of pathogenic enteric
bacteria, but not
enteric viruses or
protozoa.
Spores show extreme
environmental
persistence. Capable
of surviving longer
than waterborne
pathogens.
Insufficient data on
survival compared
with pathogens. DNA
markers show long
environmental
persistence.
Thought to be a good
model for the survival
of enteric viruses.
Phage types show
variable
environmental
persistence.
Thought to be good
models for the
survival of enteric
viruses.
Applicable to all
types of water
(fresh, estuarine
and marine).
Yes. Shorter survival
time demonstrated in
marine waters.
Yes. Demonstrates
similar survival rates
in fresh and marine
waters.
Yes. Detection
demonstrated in fresh
and marine waters.
Yes. Detection
demonstrated in fresh
and marine waters.
Yes. Detection
demonstrated in fresh
and marine waters.
Yes. Detection
demonstrated in fresh
and marine waters.
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Indicator organism
Bacteroides spp.
(genetic markers)
F+ RNA coliphages
B. fragilis phages
Insufficient data.
DNA markers show
long persistence and
may not be indicative
of recent faecal
contamination.
Yes. No significant
non-faecal sources
recognized.
Yes. No significant
non-faecal sources
recognized.
No. Dependent upon
source.
No. Dependent upon
source.
No. Dependent upon
source.
No. Dependent upon
source.
Yes. Body of
epidemiological
evidence
demonstrating
indicator shows best
correlation with health
outcomes for marine
waters and good
correlation for fresh
waters.
No. Strong correlation
with illness not
demonstrated in
epidemiological
investigations.
Insufficient data.
No. Strong correlation
with illness not
demonstrated in
epidemiological
investigations.
Insufficient data.
Yes. Culture-based
methods inexpensive,
easy to perform,
relatively rapid (24
hours), specific and
sensitive.
Yes. Culture-based
methods inexpensive,
easy to perform,
relatively rapid (24
hours), specific and
sensitive.
Yes. Culture-based
methods inexpensive,
easy to perform,
relatively rapid
(24 hours), specific
and sensitive.
No. Molecular
methods of detection
rapid, but technically
challenging and
expensive. Sensitivity
also an issue.
No. Expensive and
labour-intensive
recovery methods.
No. Complex
methodology.
Primary indicator of
faecal contamination.
Primary indicator of
faecal contamination.
Pathogen indicator;
secondary indicator of
faecal contamination.
Secondary indicator
of faecal
contamination.
Pathogen indicator;
secondary indicator of
faecal contamination.
Pathogen indicator;
secondary indicator of
faecal contamination.
Criteria
E. coli
Enterococci
C. perfringens
Absent from nonpolluted waters
and exclusively
associated with
animal and
human faeces.
Generally regarded as
true. Some evidence
of survival in organicrich environments not
associated with faecal
contamination.
Generally regarded as
true. Some evidence
of survival in organicrich environments not
associated with faecal
contamination.
No. Spores capable of
persisting in soils and
aquatic sediments.
Density directly
correlated with
the degree of
faecal
contamination.
Generally regarded as
true.
Generally regarded as
true.
Density
quantitatively
related to
swimmerassociated
illnesses.
Yes. Body of
epidemiological
evidence
demonstrating
indicator shows best
correlation with health
outcomes for fresh
waters.
Detection and
enumeration
methods rapid,
easy to perform,
inexpensive,
specific and
sensitive.
Currently
suggested role:
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5.0
Pathogenic microorganisms
Guideline values
No guideline values can be established for waterborne pathogenic microorganisms in
recreational waters. Testing for their presence in waters used for recreation should be performed
only when there is epidemiological or other evidence to suggest that this is necessary.
Background
There are three main types of pathogenic microorganisms that can be found in recreational
waters: bacteria, viruses and protozoa. Many occur as a result of contamination from human or
animal wastes, whereas some are free-living microorganisms that exist naturally in the
recreational water environment.
The challenges associated with the detection of pathogenic microorganisms in recreational
waters are currently too great to recommend this practice as part of a regular monitoring
program. Surveillance should be undertaken only during special circumstances, such as during
waterborne disease outbreak investigations.
Faecal indicators such as E. coli and enterococci are the best available surrogates for predicting
the presence of enteric pathogenic microorganisms. The presence of these indicators is expected
to indicate the possible presence of these organisms. The absence of the recommended faecal
indicators, however, should not be interpreted to mean that all pathogenic microorganisms are
also absent. Although it is not possible to completely eliminate the risk of waterborne disease,
adopting a multi-barrier approach to recreational water management will help minimize the risk
of human exposure to pathogenic microorganisms (bacteria, viruses and protozoa) in recreational
waters.
Information is provided on those pathogens recognized as having significance for Canadian
recreational waters. This list is not intended to be exhaustive, and responsible authorities may
wish to provide information on other organisms to consider regional interests. Additional
information on many of these organisms can be found in the technical documents for the
Guidelines for Canadian Drinking Water Quality.
5.1
Bacterial pathogens
A number of pathogenic bacteria can potentially be found in Canadian recreational waters.
Enteric pathogenic bacteria occur in recreational waters as a result of contamination with human
or animal faecal wastes. Sources include sewage discharges, combined sewer overflows,
stormwater, malfunctioning septic waste systems and infected swimmers. A number are
recognized zoonotic pathogens, and thus faecal shedding by animals and stormwater runoff from
areas affected by the presence of animals are also important sources. Transmission occurs via the
faecal–oral route, through accidental ingestion of contaminated waters. Gastrointestinal
symptoms are the most common manifestation of illness following infection with enteric
bacterial pathogens. Some pathogens can cause illness with more serious outcomes. E. coli and
enterococci are the best available indicators for predicting the possible presence of enteric
pathogenic bacteria.
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Other pathogenic bacteria can be free-living species or can enter natural waters through means
other than faecal contamination. Transmission can occur in waters containing sufficient
quantities of the organisms, typically by inhalation or via direct contact with body surfaces. The
types of illness caused can be varied, ranging from respiratory illnesses to infections of the eyes,
ears or skin. As these organisms are not of faecal origin, faecal indicators are not expected to
correlate well with the presence of these bacteria. Currently, there is no recognized
microbiological indicator for many of these pathogens.
5.1.1
Enteric pathogenic bacteria (Campylobacter, pathogenic E. coli, Salmonella, Shigella)
Campylobacter
Campylobacter species are Gram-negative, motile, non-spore-forming, spiral, curved or Sshaped rods. They are thermophilic (growing optimally at 42°C and incapable of growth below
30°C) and mesoaerophilic (surviving best under partially anaerobic conditions) organisms. The
genus Campylobacter is composed of 15 species; however, C. jejuni and C. coli represent the
major species of human concern in the water environment.
Campylobacter are predominantly considered to be zoonotic pathogens (Fricker, 2006). The
organisms are harboured in the intestinal tract of a wide range of domestic and wild animals,
particularly birds. Poultry is regarded as the primary avian source, and the organism has been
isolated from virtually all bird species (Fricker, 2006). It is likely that a significant proportion of
seagulls also carry these organisms (Moore et al., 2002; Pond, 2005). Cattle, sheep and pigs are
also considered to be reservoirs.
The exact mechanisms of Campylobacter virulence are incompletely understood. Attachment to
and invasion of the human intestinal tract are important factors in causing disease. C. jejuni is
reported to be capable of producing a cholera-like enterotoxin that is thought to illicit the
production of a profuse, watery diarrhoea in ill individuals.
Symptoms of Campylobacter enteritis include a profuse, watery diarrhoea (with or without blood
and/or faecal leukocytes), cramps, abdominal pain, chills and fever. The average incubation
period is 2–3 days but can span from 1 to 8 days (Percival et al., 2004). Illness is typically selflimited, requiring 3–7 days for recovery. Estimates of the number of cells generally required to
be ingested to lead to infection have ranged from as many as 104 organisms to as few as 500–
1000 cells (Percival et al., 2004; Pond, 2005). With pathogens in general, it is theorized that a
single organism is sufficient to cause human infection. Epidemiological studies have shown,
however, that the dose required is usually greater.
Certain complications (Guillain-Barré syndrome, Reiter’s syndrome, appendicitis, carditis and
meningitis) have been associated with Campylobacter enteritis; however, these are considered
rare. Fatalities from Campylobacter infection are uncommon and have been predominantly
restricted to infants, the elderly or patients afflicted with other underlying illness (Pond, 2005).
Despite the fact that Campylobacter spp. can be fairly widely isolated from surface waters, there
have been virtually no recorded instances of Campylobacter-associated illness as a result of
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Guidelines for Canadian Recreational Water Quality: Third Edition
recreational water activity in North America. The U.S. CDC reported that Campylobacter spp.
were not implicated as the causative agents in any of the recreational water outbreaks of
gastroenteritis to have been documented in the United States for the period 1992–2002 (Craun
et al., 2005). Similarly, no outbreaks of campylobacteriosis have been recorded in Canadian
recreational waters.
Pathogenic E. coli
E. coli are Gram-negative, motile, facultatively anaerobic, non-spore-forming rods that are
natural inhabitants of the intestinal tract of humans and animals. The vast majority of the E. coli
isolates are harmless; however, there are several serotypes or strains that possess virulence
factors enabling them to act as human pathogens. Pathogenic enteric strains can be separated into
six groupings according to their serological or virulence characteristics: enterohaemorrhagic
(EHEC), enterotoxigenic (ETEC), enteroinvasive (EIEC), enteropathogenic (EPEC),
enteroaggregative (EAEC) and diffuse adherent (DAEC). Human sewage is the principal source
of all of the major pathogenic E. coli groups, with the exception of EHEC. Cattle are considered
the primary reservoir for EHEC, although human wastes also remain an important source.
Of the groupings, it is the EHEC group—(also commonly referred to as verocytotoxigenic E. coli
[VTEC] or Shiga-toxin producing E. coli [STEC])—that is of greatest importance to recreational
waters. E. coli O157:H7 is the most significant serotype of this group, and has been identified as
the causative agent in a number of recreational water outbreaks (Craun et al., 2005).
An important virulence factor for EHEC is this ability to produce Shiga-like toxins, similar to
those produced by Shigella dysenteriae. EHEC infection causes haemorrhagic colitis, marked by
grossly bloody diarrhoea, severe cramping and abdominal pain with a general lack of fever. The
incubation period for the disease ranges from 1 to 8 days (Percival et al., 2004; Pond, 2005), with
the duration of infection lasting from 1 to 12 days (Percival et al., 2004). Persons only suffering
from diarrhoea typically experience a full recovery (Pond, 2005). An estimated 2–8% of all cases
progress to what is known as haemolytic uraemic syndrome, or HUS—a life-threatening
condition involving large-scale destruction of red blood cells and kidney failure. Children, the
elderly and immunocompromised persons are at increased risk for developing HUS.
The number of EHEC cells needed to be ingested in order to lead to infection is considered to be
very low. General estimates regarding the infectivity of E. coli O157:H7 suggest that ingestion of
fewer than 100 cells may be all that is necessary (Percival et al., 2004), and that as few as 50 or
even 5 organisms may be sufficient (Pond, 2005).
According to surveillance data published by the U.S. CDC for the period 1992–2002, EHEC
were associated with 25% (16 of 64) of the total number of outbreaks of gastrointestinal illness
reported for natural waters (Moore et al., 1993; Kramer et al., 1996; Levy et al., 1998; Barwick
et al., 2000; Lee et al., 2002; Yoder et al., 2004). E. coli O157:H7 was the serotype implicated in
14 of 16 of those outbreaks. The remaining two outbreaks involved E. coli serotypes O121:H19
and O26:NM.
In August 2001, an outbreak of E. coli O157:H7-associated illness involving four children was
linked to swimming at a public beach in Montreal (Bruneau et al., 2004). This was the first
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Guidelines for Canadian Recreational Water Quality: Third Edition
reported incident of E. coli O157:H7 to be associated with recreational water activity in Canada.
Weekly water samples collected around the time of the outbreak were shown to be within the
recreational water quality limits specified by the Province of Quebec. It was suggested that a
high swimmer population and the shallow depth of water encountered in the swimming area
contributed to the transmission of the organisms. To date, there have been no reported fatalities
resulting from infection with pathogenic E. coli acquired through recreational contact with
natural waters in either the United States or Canada.
Salmonella
Salmonella are members of the family Enterobacteriaceae. They are Gram-negative, facultatively
anaerobic, motile, non-spore-forming rods. The taxonomy of the genus Salmonella is quite
complex. Currently, there are over 2500 known Salmonella serotypes (or serovars) (Lightfoot,
2004). The genus is officially thought to comprise two species: S. enterica and S. bongori
(Percival et al., 2004). S. enterica can be further subdivided into six subspecies (S. enterica
subsp.): enterica, salmae, arizonae, diarizonae, houtenae and indica. The majority of the
serotypes encountered in cases of human gastroenteritis belong to the subspecies S. enterica
subsp. enterica (Lightfoot, 2004). Owing to the complexity of the nomenclature, by convention,
when referring to Salmonella, the serotype is adopted in place of the species name. Thus,
S. Enteritidis takes the place of S. enterica subsp. enterica serovar enteritidis.
With the exception of the typhoidal species (S. Typhi, S. Paratyphi), Salmonella are considered
zoonotic pathogens. Reservoirs for non-typhoidal Salmonella species include poultry, swine,
birds, cattle, rodents, tortoises and turtles, dogs and cats (Percival et al., 2004). Humans
recovering from illness can also provide a source of Salmonella, and asymptomatic infections
among humans are also possible. In contrast, humans are considered the primary source of S.
Typhi and S. Paratyphi. Occurrence of these isolates in animal hosts or in the natural
environment is rare, particularly in Canada.
Gastroenteritis represents by far the most commonly encountered type of Salmonella-associated
illness. The chief symptoms are mild to severe diarrhoea, nausea and vomiting. Symptoms
usually appear between 12 and 48 hours from the time of infection, but this time lag may be
reduced in cases where large quantities of cells have been consumed (Percival et al., 2004).
Illness is generally mild and self-limiting, lasting 2–5 days on average. Reports on the infectivity
of Salmonella have suggested that the median dose for the non-typhoidal species may be as low
as 1000 cells, and possibly below 10 cells (Hunter, 1997; Pond, 2005).
Enteric fever (typhoid or paratyphoid fever) is a more severe and often fatal form of
salmonellosis caused by S. Typhi and S. Paratyphi. Symptoms of the illness are prolonged fever,
diarrhoea and abdominal pain; progression to septicaemia may also occur. Waterborne outbreaks
of enteric fever are more prevalent in developing countries where crowded living conditions and
poor hygienic practices exist and are often associated with improperly treated drinking water
supplies. Cases are rare in North America.
Septicaemia represents the condition in which the infecting bacteria have invaded the
bloodstream and is accompanied by visible symptoms, such as high, remittent fever. Fatal
damage to the liver, spleen, respiratory or neurological functions may occur when the organism
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has spread to these organs or systems. Septicaemia from non-typhoidal species is uncommon
(Pond, 2005).
Although Salmonella can be fairly widely isolated from surface waters, there have been no
recorded outbreaks from Salmonella as a result of recreational water activity in North America.
U.S. CDC surveillance data for the years 1992–2002 indicated that Salmonella was not cited as a
causative agent for any of the waterborne outbreaks of gastroenteritis reported over that period.
Surveillance in Canada has been somewhat limited; however, similarly, there have been no
documented outbreaks from Salmonella in Canadian recreational waters.
Shigella
Shigella species are members of the family Enterobacteriaceae and, as such, possess many of the
same characteristics as E. coli. Both are Gram-negative, facultatively anaerobic, non-sporeforming rods. Unlike the majority of E. coli isolates, however, Shigella species do possess
certain attributes that make them important human pathogens. The genus Shigella is composed of
four species: S. sonnei (1 serotype), S. flexneri (6 serotypes), S. boydii (15 serotypes) and
S. dysenteriae (10 known serotypes). Two species, S. sonnei and S. flexneri, account for the vast
majority of Shigella-associated illness in North America (CDC, 2005a). Other Shigella species
are uncommon, but remain important causes of disease in developing countries (CDC, 2005a).
Humans are considered to be the only important reservoir for Shigella (Percival et al., 2004).
Non-infected individuals are not expected to harbour the organism. Recovering individuals may
continue to shed significant numbers of the bacterium for several weeks after symptoms have
resolved, and asymptomatic carriage of Shigella is also possible. Municipal sewage discharges
present an obvious source of Shigella; however, faecal shedding by infected swimmers may be
the most significant source of the organism in recreational waters (Kramer et al., 1996; Levy et
al., 1998).
Shigella causes what had been historically referred to as bacillary dysentery (invasion of the
intestinal tract, causing frequent passage of stools containing blood and mucus). Illness
(shigellosis) is caused by invasion and colonization of the intestinal tract, which leads to
inflammation and the destruction of intestinal epithelial cells. Shigella species are also capable of
producing a heat-labile enterotoxin; however, its role is not completely understood (Percival et
al., 2004).
Shigellosis is characterized by watery or bloody diarrhoea, abdominal pain and fever. Symptoms
often present within 1–3 days post-infection, but may be evident in as little as 12 hours. The
severity of illness is strongly dependent on the virulence of the individual species or strain.
S. sonnei infections are thought to be both of shorter duration and somewhat milder than those
caused by S. flexneri (Percival et al., 2004). In many developing countries, S. dysenteriae is
known to be responsible for causing severe epidemics. S. dysenteriae serotype 1 is also known to
be capable of producing what is referred to as Shiga toxin—a unique toxin (separate from the
general Shigella enterotoxin) that can be extremely damaging to human intestinal and kidney
endothelial cells. For S. sonnei or S. flexneri, an inoculum of 100 cells may be sufficient to
produce infection, whereas S. dysenteriae may require the ingestion of as few as 10 cells to
initiate disease (Pond, 2005).
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In North America, most cases of shigellosis are mild and self-limiting. Infection typically does
not spread beyond the intestinal tract. Complications such as Reiter’s syndrome and HUS
(following infection with S. dysenteriae serotype 1) have been reported, but are uncommon.
Similarly, mortality is rare, but higher incidences may be observed among the elderly and in
undernourished children (Pond, 2005).
According to CDC surveillance data, Shigella accounted for approximately 22% (14 of 64) of the
total number of outbreaks of gastrointestinal illness reported for natural recreational waters in the
United States over the period 1992–2002 (Moore et al., 1993; Kramer et al., 1996; Levy et al.,
1998; Barwick et al., 2000; Lee et al., 2002; Yoder et al., 2004). S. sonnei was implicated as the
causative agent in all but one of these incidences. Recreational lakes provided the setting for the
majority of the outbreaks reported, with poor water circulation frequently cited as a contributing
factor. It was suspected that faecal contamination by other swimmers was the cause in the
majority of these instances. Given the high infectivity reported for Shigella, it is thought that
accidental swallowing of water containing relatively low concentrations of these organisms may
be sufficient to cause illness.
Although surveillance has been somewhat limited, to date there have been no reported incidences
of Shigella-associated illness as a result of recreational water activity in Canadian waters.
5.1.2
Free-living pathogenic bacteria (Aeromonas, Legionella, Pseudomonas, Mycobacterium)
Aeromonas
Aeromonas are Gram-negative, facultatively anaerobic, variably motile, oxidase-positive, rodshaped to coccoid-like bacteria. They are thought to share many morphological and biochemical
characteristics with members of the Enterobacteriaceae family, which includes E. coli.
Currently, the genus Aeromonas is thought to consist of 17 unique genospecies and 14 unique
phenospecies (Moyer, 2006; U.S. EPA, 2006a). The genus may be subdivided into two major
groups: the motile mesophilic species, which grow at temperatures between 15 and 38°C and
have been associated with human infections, and the non-motile psychrophilic species, which
can grow at temperatures below 15°C and are fish pathogens. At present, six species (A.
hydrophila, A. caviae, A. sobria, A. veronii, A jandaei, A. trota and A. schubertii) are recognized
as being pathogenic to humans (U.S. EPA, 2006a).
Aeromonas species are natural inhabitants of the aquatic environment. They are frequently
found in fresh, marine and estuarine waters, sediments, and sewage and wastewater effluents.
Aeromonads are not considered to be present in significant numbers in the faeces of healthy
individuals. However, a certain percentage of individuals may carry the organisms in their
intestinal tract without showing outward signs of illness.
Aeromonas are recognized animal pathogens and have been isolated from the intestinal tract of
a number of animal species, including fish, reptiles, amphibians, birds and domestic livestock,
with and without evidence of illness. Occurrence of the organism in recreational waters is not
dependent on faecal pollution; however, the organisms are present in high numbers in sewage
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and thus can be detected at significant levels in sewage-contaminated waters. Aeromonads can
grow to relatively high densities in eutrophic waters (Moyer, 2006).
Aeromonas is most often associated with serious wound infections in recreational water users.
Infection typically requires the existence of some sort of skin trauma, such as an open wound or
following some sort of penetrating injury. Wound infections are characterized by pain, swelling,
redness and fluid accumulation around the infected area. Cellulitis (severe inflammation) is
frequently observed with such infections, and septicaemia is also considered a fairly common
outcome (Percival et al., 2004). Other, rarer complications include necrotizing fasciitis,
meningitis, pneumonia, peritonitis and endocarditis (Percival et al., 2004).
Several Aeromonas species (A. hydrophila, A. veronii and A. caviae) have been associated with
human gastrointestinal illness. Gastrointestinal illness in humans exposed to contaminated water
supplies has been only occasionally reported. Illness is typically mild and self-limiting, although
certain strains are reported to be capable of causing a dysentery- or cholera-like illness, marked
by severe abdominal cramps, vomiting, diarrhoea (including bloody stools) and fever.
The mechanisms through which Aeromonas is able to cause human illness are not completely
understood. The organisms possess a host of virulence factors considered important for infection,
colonization and evading the host’s immune response. These include both cell-associated
mechanisms (pili, flagella, outer membrane proteins, lipopolysaccharides and capsules) and
extracellular products (toxins, proteases, haemolysins, adhesions and various hydrolytic
enzymes) (U.S. EPA, 2006a).
Marino et al. (1995) reported a positive correlation between A. hydrophila concentrations and
skin infections at two swimming beaches in Malaga, Spain. Currently, no evidence has been
provided linking Aeromonas concentrations and the risk of acquiring swimming-associated
gastroenteritis.
Despite their widespread occurrence, in North America there have been no reported outbreaks
of Aeromonas-associated illness as a result of recreational water activities. Superficial infections
with Aeromonas are thought to be relatively common; however, Aeromonas infections are not
considered reportable illnesses. Thus, an estimate of the likely incidence of Aeromonas
infections due to recreational water exposures in Canadian waters is not available.
Legionella
Legionella are Gram-negative, thermotolerant, motile, short, irregularly shaped bacteria that have
strict nutrient requirements when grown on laboratory media. Over 40 species of the genus have
been recognized. L. pneumophila (serotype 1) is the species most frequently associated with
human disease (legionellosis). It is suspected that all Legionella species may be capable of
causing illness, and approximately half of the identified species have been implicated in human
disease (Hall, 2006). Other Legionella species frequently recovered from the environment
include L. bozemanii, L. longbeachae, L. dumoffii and L. gormanii.
Legionella are naturally occurring aquatic organisms. They can be isolated from a wide range of
freshwater habitats, including soils, lakes, rivers and natural thermal pools at temperatures as
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high as 60°C. Marine environments typically do not provide appropriate growth conditions for
these organisms. Free-living freshwater protozoa such as Naegleria or Acanthamoeba are natural
hosts for these organisms. Although Legionella are thought to be fairly resistant to environmental
stresses, survival within protozoan hosts is thought to provide an additional, significant measure
of protection.
Legionella is typically encountered in low numbers in the aquatic environment, but can reach
higher concentrations in sources associated with human-made water supplies, such as cooling
towers, air conditioning condensers, humidifiers, hot water tanks, shower heads and whirlpool
spas (Percival et al., 2004). Hot springs or other hydrothermal spas are particularly well suited
to the survival of Legionella as a result of the elevated water temperatures.
Dose–response experiments with animals have suggested that high doses of Legionella
(approximately 107 cells) are required to initiate infection (O’Brien and Bhopal, 1993). In
contrast, published reviews of the organism have suggested that the median dose is as low as a
few organisms (Percival et al., 2004; Pond, 2005). Aerosol carriage of protozoa that have been
heavily parasitized by Legionella may be one means of increasing the number of organisms
available to initiate infection (Percival et al., 2004).
Legionella are important agents of respiratory disease in humans. Legionellosis comprises two
forms of illness: Legionnaire’s disease and Pontiac fever (Pond, 2005). Legionnaire’s disease is
a more severe and sometimes fatal form of respiratory illness. The disease is defined by the
clinical diagnosis of pneumonia, accompanied by microbiological evidence of infection with
L. pneumophila or other Legionella species (Pond, 2005). Other symptoms can include fatigue,
fever, headache, muscle and/or abdominal pain, jaundice and mental confusion. The incubation
period is between 3 and 6 days, and recovery is slow, lasting weeks up to several months.
Fatalities can result, largely due to respiratory failure. The mortality rates from communityacquired infections are estimated to be in the range of 5–20% (Pond, 2005).
Pontiac fever is a relatively mild, influenza-like illness, defined by non-pneumonic
respiratory illness with microbiological evidence of Legionella infection. The disease has
a shorter incubation period (1–2 days), and illness is considered non-fatal, with infected
individuals requiring approximately 2–5 days for recovery. It is estimated that Pontiac
fever occurs 2–100 times more frequently than Legionnaire’s disease (Hall, 2006).
Groups considered most sensitive to Legionella infection include the elderly,
immunocompromised individuals, persons with heart or lung disease and individuals who
are excessive smokers or consume excess amounts of alcohol.
Despite the fact that Legionella species are thought to be ubiquitous in environmental waters, no
recorded outbreaks of legionellosis have been reported in Canada or the United States as a result
of recreational activity in natural waters. Any reported outbreaks of legionellosis associated with
human recreational water contact have been restricted to treated water facilities, such as hot tubs
and spas (Moore et al., 1993; Kramer et al., 1996; Levy et al., 1998; Barwick et al., 2000; Lee et
al., 2002; Yoder et al., 2004).
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Mycobacterium
Mycobacterium species are aerobic, non-motile, non-spore-forming, rod- to coccoid-shaped
bacteria. The organisms are considered to have a Gram-positive cell structure. However, the
mycobacterial cell wall contains high levels of mycolic acid—complex lipids that give the cell
surface a waxy, hydrophobic character that resists Gram staining. Positive staining with acid-fast
staining techniques is diagnostic for mycobacteria; thus, the organisms are more commonly
referred to as “acid-fast.”
The pathogenic mycobacteria encountered in recreational waters are environmental species.
They are generally referred to as “atypical” or “non-tuberculous” mycobacteria to distinguish
them from M. tuberculosis (tuberculosis) and M. leprae (leprosy). Neither M. tuberculosis nor
M. leprae is found in the environment. Consequently, they are not a concern for recreational
waters. At least 16 different waterborne species have demonstrated the capability for causing
infection in humans (Pond, 2005). The species most commonly discussed as having relevance for
recreational water exposures are the members of the Mycobacterium avium complex (M. avium
and M. intracellulare), which are known to cause respiratory illness; and M. marinum and
M. kansaii, which can cause skin infections.
Environmental mycobacteria are considered ubiquitous in natural waters. They can be found in
virtually every medium, including soils, wastewater, lakes, rivers, ponds, streams, groundwater
and treated water supplies. Few mycobacteria are encountered in marine waters (Pond, 2005;
LeChevallier, 2006). The organisms can survive over a wide range of temperatures, extending
from below 0°C to above 50°C. Mycobacteria are not known to be present in high numbers in
faeces, and sewage is not considered to be a significant source (Falkinham, 2002). M. avium
complex members are capable of survival and growth within certain species of phagocytic
protozoa, specifically members of the genus Acanthamoeba.
In recreational water environments, transmission can occur through contact with waters
containing sufficient quantities of the organisms. The main routes of infection are via inhalation
of mycobacteria contained within aerosols and through direct water contact with abraded skin.
There is little evidence of person-to-person transmission. Environmental mycobacteria
are primarily considered to be opportunistic pathogens, as illness is more commonly observed in
individuals with some underlying condition that predisposes them to infection (abraded or
traumatized skin; or having a weakened or compromised immune system). Exposures to
environmental mycobacteria have been most strongly linked to swimming pool and hot tub
use, resulting in cases of skin and soft tissue infections and hypersensitivity pneumonitis
(inflammation of the lungs). Recreational contact with natural waters is not considered to be
a significant risk factor for acquiring mycobacterial illness.
Although environmental mycobacteria are considered ubiquitous in most types of water, to date
there have been no recorded outbreaks of mycobacterial-associated illness through contact with
natural recreational waters in either Canada or the United States. The risk of healthy individuals
acquiring a mycobacterial infection as a result of recreational activity in natural waters is
considered extremely low.
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Pseudomonas aeruginosa
Pseudomonads are Gram-negative, motile, oxidase-positive, non-spore-forming, rod-shaped
bacteria. Over 100 species are currently recognized as belonging to the genus Pseudomonas
(Hunter, 1997). P. aeruginosa represents the most significant species of human concern.
P. aeruginosa is widely distributed in the aquatic environment and can be frequently isolated
from fresh water, seawater and soils (Hunter, 1997). The organism has minimal growth
requirements and is able to proliferate in waters of low nutrient content. P. aeruginosa is
infrequently isolated from human faeces (Geldreich, 2006). The organism can be recovered from
sewage and stormwater (as these contain a mixture of domestic wastes) and from industrial
discharges such as food processing and pulp and paper wastes. Swimmers themselves also
present another possible source of P. aeruginosa.
Transmission of P. aeruginosa in recreational waters occurs through direct body contact with
waters containing sufficient quantities of the organism. Ingestion is not considered to be a
significant route of infection.
P. aeruginosa can be responsible for causing skin rashes and eye and ear infections among
recreational water users. Infection rarely occurs among healthy individuals unless some
condition exists that might predispose them to infection (having a history of ear infections, or
reporting frequent immersions) (Hunter, 1997). Ear infections occur when P. aeruginosa is able
to enter into and colonize the outer ear canal. Within a few days of swimming, the ear may
become itchy and painful, and discharges of pus may be observed. Skin irritations (dermatitis)
present as a red, itchy rash, occurring roughly 18–24 hours after water contact. Infection can
progress to folliculitis (inflammation of the hair follicles of the skin), which is marked by an
increased tenderness of the infected area and the presence of pus-filled blisters or pimples that
surround the hair follicles.
Several epidemiological studies have demonstrated the existence of a link between Pseudomonas
in natural waters and the incidence of eye and skin infections among swimmers (Seyfried and
Cook, 1984; Springer and Shapiro, 1985; Ferley et al., 1989; Marino et al., 1995; van Asperen et
al., 1995). Reported outbreaks of Pseudomonas dermatitis have virtually all been associated with
treated water venues such as hot tubs, swimming pools or hotel whirlpool or spa baths (Moore et
al., 1993; Kramer et al., 1996; Levy et al., 1998; Barwick et al., 2000; Lee et al., 2002; Yoder et
al., 2004). The incidence of P. aeruginosa infections from contact with natural recreational
waters is not known, as illnesses are usually of mild severity and typically not recorded.
5.1.3
Other pathogenic bacteria (Leptospira, Staphylococcus)
Leptospira
Leptospira are spirochetes—spirally coiled or corkscrew-shaped bacteria. They are Gramnegative staining, aerobic, long, thin and motile organisms. Initially, the genus Leptospira
consisted of two species, the pathogenic L. interrogans and the free-living L. biflexa (WHO,
2003b). At present, 12 Leptospira species are recognized, and over 200 pathogenic serotypes
have been described, in which the more severe forms are attributed to serovars of L. interrogans
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(Pond, 2005). By convention, the serotype name is often adopted as the species name when
referring to specific strains.
Leptospira species can be either pathogenic or free-living. They are encountered worldwide and
are predominantly associated with freshwater environments. The pathogenic leptospires are
important zoonotic pathogens that are carried in the renal tract (kidney) of infected animal
carriers and excreted in the urine. Small rodents, such as rats, mice and voles, are considered the
most important source of pathogenic Leptospira. The organisms can also be spread by domestic
animals, such as cattle, pigs, dogs and cats, sheep, goats and horses (WHO, 2003b; CDC,
2005b). Heavy rainfall is thought to facilitate the spread of the organisms, as runoff from
contaminated soils can affect surface waters (Pond, 2005).
Human infection can occur following direct contact with the urine of infected animals or
indirectly though contact with contaminated water, soil or mud. Leptospires gain access through
cuts or abrasions in the skin or via passage through the mucous membranes of the eyes, nose and
mouth. Ingestion of contaminated water and inhalation of leptospires carried in aerosols are also
possible routes of infection. It is thought that ingestion of as few as 1–10 organisms can be
sufficient to lead to human illness (Pond, 2005). Recreational water activity is perhaps the most
significant source of exposure for acquiring illness, although swimming-associated outbreaks are
considered extremely rare (Pond, 2005).
Illness following infection with Leptospira can range in severity from a mild, influenza-like
illness to more severe, and possibly fatal, disease. Early symptoms of illness are fever, chills,
headache, muscle pains, vomiting and reddening of the eyes (Public Health Agency of Canada,
2004). Recovery from mild illness is usually complete, but can be lengthy, in some cases
requiring months to years (WHO, 2003b). If left untreated, the disease can progress to more
serious illness. Severe cases of leptospirosis can be fatal, with death occurring as a result of
kidney failure, cardiorespiratory failure or extensive haemorrhaging. The reasons for the
differences in the severity of infection are not fully understood; however, it is believed that each
pathogenic serovar possesses the capacity to cause either mild or severe disease (WHO, 2003b).
Illness can be difficult to diagnose, as it may be mistaken for other infections or illnesses that
produce similar symptoms. Similarly, mild forms of the illness may not always be reported.
Leptospirosis is considered to be a greater concern among developing countries and in tropical
climates. There have been three reported outbreaks of leptospirosis in recreational waters in the
United States over the period 1991–2002 (Moore et al., 1993; Barwick et al., 2000; Lee et al.,
2002).
Reports of increases in the number of observed cases of leptospirosis in developed countries
suggest that Leptospira may represent an important re-emerging pathogen (CRC, 2004; Meites et
al., 2004). Currently, the prevalence of Leptospira in Canadian waters is not known. To date,
there have been no documented incidences of Leptospira infection from recreational water
activity in Canadian waters.
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Staphylococcus aureus
Members of the genus Staphylococcus are Gram-positive, catalase-positive, non-motile cocci.
S. aureus is considered the major pathogen of the genus and is the species of most significance
for recreational water users.
S. aureus is not considered to be a natural inhabitant of environmental waters. The major
reservoirs for this organism are the skin, nose, ears and mucous membranes of warm-blooded
animals. The presence of S. aureus in recreational waters is predominantly due to releases of the
organism from the mouths, noses and throats of swimmers and from discharges from existing
infections. The organism can be isolated from human faeces; however, occurrence is thought to
be variable (Percival et al., 2004). Sewage and stormwater are additional sources of the
organism.
Transmission of S. aureus in recreational waters occurs via direct contact with waters containing
sufficient quantities of the organism. Infection occurs through cuts or scratches on the skin or, to
a lesser extent, through contact with the eyes and ears. Person-to-person spread of the organism
is also possible. Ingestion is not considered to be a significant route of exposure. The organism
produces a wide array of extracellular toxins, exoenzymes and adherence factors that are used
during colonization and infection and for evading host immune defences (Percival et al., 2004).
Concentrations of a few hundred cells per millilitre may be sufficient to initiate infection in
injured or distressed skin (Percival et al., 2004).
S. aureus is predominantly associated with skin infections in recreational water users. Common
infections include infected cuts and scratches, boils, pustules, dermatitis, folliculitis and impetigo
(WHO, 2006). Infections are most often pus-forming, with symptoms often not becoming
apparent until 48 hours after contact. Other illnesses to which the organism has been linked
include eye infections, otitis externa and urinary tract infections (WHO, 2006).
Epidemiological investigations have demonstrated evidence of possible connections between the
presence of staphylococci in recreational waters and swimmer illness (Calderon et al., 1991;
Charoenca and Fujioka, 1995). However, to date, there has been no conclusive evidence relating
the frequency of illness to the concentration of S. aureus in recreational waters. In certain
instances, testing may provide additional information—for example, in assessing the effects of
high swimmer densities on water quality and the potential implications for the possible personto-person transfer of pathogens.
5.2
Viral pathogens
Viruses are submicroscopic organisms, much smaller in size than bacteria. They are simply
constructed, consisting of a nucleic acid core composed of either RNA or DNA, and surrounded
by an external protein shell called a capsid. The nucleic acid encodes for viral structural proteins
and enzymes necessary for replication, whereas the capsid protects the viral unit from
environmental stresses. Some viruses (enveloped viruses) may also possess a lipoprotein
envelope surrounding the capsid. Non-enveloped viruses lack this external layer. Viruses are
obligate intracellular parasites and must infect a host cell in order to replicate. As a result, they
are incapable of replicating outside of their host environment.
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The pathogenic viruses of concern for recreational waters are enteric viruses—viruses that infect
the human gastrointestinal tract and are shed in human faeces. These viruses are considered to
have a narrow host range, meaning that in general, enteric viruses that infect animals do not
infect humans, and vice versa. Transfer to humans in recreational waters occurs via the faecal–
oral route through the accidental ingestion of contaminated waters. Some viruses, like the
adenoviruses, have additional routes of infection, such as via inhalation or through contact with
mucosal membranes of the eyes. Enteric viruses cause a wide variety of human health effects,
which can range in severity from mild to severe. Gastrointestinal symptoms (nausea, vomiting,
diarrhoea) are the most commonly encountered symptoms of viral illness. Some virus infections
can result in more serious health outcomes, although these are considered to be much rarer.
5.2.1 Enteric viruses
There are over 100 types of viruses that can be excreted in faeces and thus can potentially be
transmitted to recreational waters. The viruses most commonly associated with waterborne
illness include adenoviruses, astroviruses, enteroviruses (polioviruses, coxsackieviruses and
echoviruses), noroviruses, rotaviruses and the Hepatitis A virus.
Enteroviruses
Enteroviruses are a large group of small (20–30 nm), non-enveloped RNA viruses belonging
to the family Picornaviridae. Members of this group include polioviruses, coxsackieviruses,
echoviruses and several yet unclassified enteroviruses. Many enterovirus infections are
asymptomatic. The symptoms and severity of illness vary considerably among the individual
virus types and serotypes. The most commonly observed health effects are vomiting, diarrhoea,
febrile flu-like symptoms, malaise, respiratory disease, headache and muscle ache (Percival et
al., 2004). More serious outcomes have been associated with individual virus groups, including
myocarditis (coxsackievirus), aseptic meningitis (coxsackievirus, poliovirus), encephalitis
(coxsackievirus, echovirus) and poliomyelitis (poliovirus), although these are not considered
to be common.
Noroviruses
The term “norovirus” has been designated as the official name for the group of viruses formerly
known as Norwalk viruses, Norwalk-like viruses or “small, round, structured viruses” (SRSV).
Noroviruses are small (27–30 nm), non-enveloped RNA viruses. Norovirus infection is
considered to be the leading cause of viral gastroenteritis outbreaks (from all sources) in the
United States and the United Kingdom (Percival et al., 2004). The primary symptoms of illness
are diarrhoea, vomiting, headache and muscle ache. The onset of projectile vomiting is
considered a characteristic trait of norovirus infection. Asymptomatic infections with norovirus
are rare. In healthy adults, illness is self-limiting, rarely progressing to more serious concerns
(e.g., dehydration). Infection is considered more serious among vulnerable groups such as the
elderly.
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Rotaviruses
Rotaviruses are larger (60–80 nm), non-enveloped RNA viruses. Dose–response studies have
suggested that rotaviruses are the most infective of all the enteric viruses (Gerba et al., 1996).
Rotavirus infection has been implicated as the number one cause of infantile gastroenteritis
worldwide. Although all age groups can be affected, infections among healthy adults are often
asymptomatic as a result of the immunity acquired during childhood (Percival et al., 2004).
Diarrhoea constitutes the predominant symptom of illness, which can become life-threatening
should the resulting dehydration and electrolyte imbalance become severe. Groups considered
vulnerable for severe disease- and illness-induced mortality include young children,
immunocompromised individuals and the elderly.
Adenoviruses
Adenoviruses are also larger by comparison (70–100 nm) and are non-enveloped DNA viruses.
Over 49 individual serotypes have been identified as being capable of causing human illness,
with the clinical features and severity of illness varying considerably among the individual types
(Percival et al., 2004). The majority of adenovirus serotypes cause respiratory illness, which
presents with pharyngitis and cough and cold-like symptoms. Conjunctivitis can also occur as a
result of infection of the eye. Gastrointestinal illness, caused exclusively by serotypes 40 and 41,
is also a frequently reported outcome. Adenovirus is thought to be second only to rotaviruses as
a cause of childhood gastroenteritis (Crabtree et al., 1997). Asymptomatic illness is commonly
observed with gastrointestinal infection, as it is believed that the immunity conferred during
early childhood is lifelong (Percival et al., 2004).
Hepatitis A virus
The Hepatitis A virus (HAV) is a small (25–28 nm), non-enveloped RNA virus whose major
target organ is the liver. The majority of HAV infections are asymptomatic. Illness is most
frequently reported among adults. Symptoms include malaise and fever, followed by nausea,
vomiting, abdominal pain and, ultimately, jaundice. Infection is typically self-limiting.
Astroviruses
Astroviruses are small (28–30 nm), non-enveloped RNA viruses. Of the viral agents known to
cause enteric illness, the significance of astroviruses as a cause of waterborne illness is perhaps
the least well characterized (Percival et al., 2004). Illness in infected individuals appears similar
to rotaviral illness, although markedly less severe.
Occurrence in the environment
Enteric viruses are shed in high numbers in the faeces of infected individuals and can reach
concentrations as high as 1010–1012 particles per gram of faeces (Gerba, 2000). Even
asymptomatic individuals (those infected, but not exhibiting symptoms of disease) are capable
of excreting large numbers of viruses.
The principal route of entry for viruses in recreational waters is via the discharge of sewagecontaminated wastes. Point sources of pollution such as municipal sewage discharges or
combined sewer overflows constitute the primary sources of sewage contamination. Non-point
sources capable of contributing to the viral loading of environmental waters include storm drains,
river discharges (which capture runoff from urban and rural areas) and malfunctioning or
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improperly designed septic waste systems. Swimmers themselves, particularly young children,
can also present a source of contamination through faecal shedding and the accidental release of
faecal material. Animal wastes, although capable of harbouring many bacterial and protozoan
pathogens, are considered to be of low risk for the transmission of viruses to humans (Cliver
and Moe, 2004; Percival et al., 2004). There have been some examples of animals serving as a
reservoir for human viruses (avian influenza virus, West Nile virus), yet to date there has been
no documented evidence of human waterborne infection having been caused by animal viruses
(Cliver and Moe, 2004).
The total viral load of sewage can be quite constant; however, the types and numbers of
individual viruses are strongly influenced by the rates of epidemic and endemic illness within the
discharging population. As a result, the viral composition of sewage can vary considerably, often
demonstrating strong seasonal trends (Krikelis et al., 1985; Tani et al., 1995; Pina et al., 1998;
Lipp et al., 2001). Published estimates of culturable viruses in raw sewage suggest that
concentrations can reach over 10 000 infectious units per litre (Reynolds et al., 1998; Payment et
al., 2001). The presence of viruses in surface waters is expected to vary regionally and is
dependent upon (among other factors) the degree and type of faecal contamination and the rates
of environmental inactivation. Detectable levels of culturable enteroviruses in surface waters in
general have ranged from 1–10/100 L to 1–200/L for more contaminated waters (Pina et al.,
1998; Reynolds et al., 1998; Payment et al., 2000; Lipp et al., 2001).
Studies have reported the detection of enteroviruses, noroviruses, rotaviruses, adenoviruses,
HAV and astroviruses in marine and fresh waters used for recreational purposes in the United
States, Europe and Canada (Payment, 1984; Puig et al., 1994; Pina et al., 1998; Griffin et al.,
1999; Chapron et al., 2000; Payment et al., 2000; Schvoerer et al., 2001; Denis-Mize et al., 2004;
Jiang and Chu, 2004; Laverick et al., 2004). The studies report varying figures in terms of virus
detection frequencies and concentrations. Differences in the analytical methods used in the
respective studies prevent direct comparisons of the results; nevertheless, the information
provided by these investigations serves to shed some light on the potential vulnerability of
recreational waters to contamination with pathogenic viruses.
Viruses are hardy microorganisms that can survive for prolonged periods once shed into the
aquatic environment. Survival is dependent upon a number of biological and environmental
factors, including the virus’s specific physical characteristics, the presence of natural microbial
predators and various water characteristics, such as temperature, pH, salinity, turbidity and
ultraviolet (UV) levels. Data on the survival of individual virus types in natural waters have been
limited. Viruses are generally regarded to be more resistant to environmental degradation than
bacteria, and experimental data suggest that some enteric viruses may demonstrate greater
resistance than some enteric protozoa (e.g., Giardia) (Johnson et al., 1997).
Epidemiology
Surveillance data on recreational water outbreaks published by the U.S. CDC indicated that for
the period 1991–2002, 13% (8 of 64) outbreaks of gastroenteritis reported in natural waters were
caused by noroviruses (Moore et al., 1993; Kramer et al., 1996; Levy et al., 1998; Barwick et al.,
2000; Lee et al., 2002; Yoder et al., 2004). In general, noroviruses were responsible for between
0 and 2 outbreaks per year, with the total number of cases per outbreak ranging from 11 to
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168 individuals. Recreational lakes constituted the setting for the majority of the outbreaks, and
a 2002 outbreak involving 44 cases at Lake Michigan State Park in Wisconsin was the first
documented recreational water outbreak at a Great Lakes beach (Yoder et al., 2004). No other
virus types were implicated in any of the other outbreaks reported from 1991 to 2002.
Outbreaks of acute gastroenteritis in which the causative agent could not be identified also
regularly constitute a significant proportion of the total number of recreational water-related
outbreaks. It is generally suspected that many of these outbreaks are of viral origin. Pathogenic
viruses are notoriously difficult to detect, and the short incubation times, range of symptoms
encountered and high frequency of illness observed among children are all consistent with viral
infections (Cabelli, 1983; Mena et al., 2003; Percival et al., 2004). According to CDC
surveillance data, outbreaks of acute gastrointestinal illness of unknown etiology accounted for
23% (14 of 64) documented outbreaks from 1991 to 2002 (Moore et al., 1993; Kramer et al.,
1996; Levy et al., 1998; Barwick et al., 2000; Lee et al., 2002; Yoder et al., 2004). Again,
freshwater lakes were identified as the most frequent setting for such outbreaks.
Several epidemiological studies have attempted to characterize the relationship between
enteroviruses in swimming water and the incidence of recreational water illness (Lightfoot,
1988; Alexander et al., 1992; Fewtrell et al., 1992; van Dijk et al., 1996; Lee et al., 1997; van
Asperen et al., 1998; Haile et al., 1999). In general, no significant relationships could be shown
between the concentration of enteric viruses and the incidence of swimmer illness. Haile et al.
(1999) did observe an increase in the reporting of a number of adverse health effects (vomiting,
fever, sore throat and highly credible gastrointestinal illness) on days on which enteroviruses
were detected in swimming waters affected by stormwater discharges.
Relationship with indicators
A number of studies have reported a lack of a relationship between the concentration of faecal
indicator bacteria and the presence of enteric viruses in recreational waters (Griffin et al., 1999;
Schvoerer et al., 2000, 2001; Jiang et al., 2001; Noble and Fuhrman, 2001; Denis-Mize et al.,
2004; Jiang and Chu, 2004; Wetz et al., 2004). Pathogenic viruses have been detected in
recreational waters at faecal indicator bacteria concentrations below the existing limits for
recreational water quality. The reverse has also been true—that waters with indicator counts well
above recreational water limits have yielded negative results for the presence of viruses. The lack
of a correlation between faecal indicators and enteric viruses is not unexpected, as faecal
indicators are present consistently in human and animal wastes and in relatively constant
numbers, whereas viruses are specific to human wastes and shedding may be intermittent and
seasonal. Viruses are also more resistant than bacteria to environmental stresses and may persist
for longer periods. Other organisms have been proposed as potential surrogate indicators (strains
of enterovirus, C. perfringens, coliphages and phages of B. fragilis); however, investigations
conducted to date have not demonstrated conclusive evidence of a link between these organisms
and the detection of viruses in contaminated surface waters (Pina et al., 1998; Griffin et al., 1999,
Lipp, 2001; Jiang and Chu, 2004).
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5.3
Protozoan pathogens
Pathogenic protozoa of importance to recreational waters include both enteric and free-living
species. Enteric protozoa are common parasites that infect the intestinal tract of humans and
other mammals. They are obligate parasites, meaning that they require the infection of a host to
replicate and are incapable of growth outside the host environment. The most important stage of
their life cycle involves the production of cysts or oocysts that are shed in large numbers in the
faeces. These cysts or oocysts are extremely resistant to environmental stresses and can survive
for long periods in the environment. Upon ingestion by a new host, the (oo)cysts undergo
excystation in the small intestine to initiate infection. These organisms can enter recreational
waters as a result of direct or indirect contact with human or animal faeces. Transmission to
humans occurs through the accidental ingestion of contaminated waters. The most common
manifestations of illness are gastrointestinal symptoms, specifically diarrhoea. E. coli and
enterococci are used to indicate faecal contamination and thus the possible presence of these
faecal enteric pathogens. Indicator absence does not necessarily indicate that enteric protozoa are
also absent.
Free-living protozoa, unlike enteric protozoa, occur naturally in recreational waters and do not
require the presence of a host organism to complete their life cycle. Transmission to humans can
occur in waters containing sufficient quantities of the organisms through mechanisms such as
inhalation or through direct contact with mucous membranes (e.g., those of the eye). The types
of illnesses caused by these organisms are varied and include infections of the central nervous
system and eye infections. As these organisms are not of faecal origin, faecal indicators are not
expected to correlate well with the presence of these protozoans. Currently, there is no
recognized microbiological indicator for these pathogens.
5.3.1 Enteric protozoa (Giardia, Cryptosporidium)
The enteric protozoa of most importance to recreational waters are Giardia and
Cryptosporidium.
Giardia
Giardia spp. are small, flagellated protozoan parasites. Species have a two-staged life cycle
consisting of a trophozoite (feeding stage) and an environmentally resistant cyst stage.
G. duodenalis (syn. lamblia, intestinalis), found in humans and a wide range of other mammals,
is the only human-infective species. Other species (G. muris, G. agilis, G. microti, G. psittaci
and G. ardea) have been reported in animals, including rodents, birds and amphibians. Molecular
characterization of G. duodenalis has demonstrated the existence of genetically distinct
genogroups (assemblages) depending on their host range. Some groups have shown occurrence
across both human and animal hosts, whereas others have been shown to be host-specific.
Human and animal faeces (especially cattle) are major sources of G. duodenalis. Other
recognized animal hosts include beavers, muskrats, dogs, sheep and horses. Many of these
animals can be infected with G. duodenalis originating from human sources (Davies and Hibler,
1979; Hewlett et al., 1982; Erlandsen et al., 1988). Epidemiological and molecular data suggest
that it is only these human-source strains that have been significantly associated with
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human illness. The pathogenicity of other, animal-specific G. duodenalis strains and Giardia
species is not fully known. As a result, it remains sound practice to consider any Giardia cysts
found in water as potentially infectious to humans.
Giardia is commonly encountered in sewage and surface waters. In general, concentrations in
wastewater are in the range of 5000–50 000 cysts/L, with surface water concentrations typically
ranging from < 1 to 100 cysts/100 L (Medema et al., 2003; Pond et al., 2004).
The exact mechanisms through which Giardia causes illness are not completely understood.
Damage to the intestinal mucosa caused by attachment and detachment of the trophozoites
contributes to the impairment of intestinal function. The severity of Giardia infection can range
from no observable symptoms to severe gastrointestinal illness requiring hospitalization. The
most common symptoms of illness include explosive, watery diarrhoea, nausea, intestinal upset,
fatigue, low-grade fever and chills.
In theory, a single cyst is sufficient to cause human infection. However, studies have shown that
the dose required to cause infection is usually greater. Human (volunteer) feeding studies have
suggested that the median dose for infection is around 50 cysts (Hibler et al., 1987), although
subjects have shown infection at doses much lower than this (Rendtorff, 1978). The time
between ingestion and the excretion of new cysts (prepatent period) ranges between 6 and
16 days. Infection is self-limiting, clearing within 1–3 weeks on average. Some patients may
remain as asymptomatic carriers, whereas in other cases individuals may experience recurrent
bouts of the disease, a phase persisting for a period of several months to a year. Persistent illness
can be treated using a number of antiparasitic drugs.
Cryptosporidium
Cryptosporidium are small, non-motile protozoan parasites. These organisms possess a complex,
multi-staged life cycle, of which the most important stage is production of the round, thickwalled oocysts. Sixteen species are currently recognized as belonging to the genus. Two
predominant genotypes have been linked to human illness: C. hominis (genotype 1), reported
only in humans, and C. parvum (genotype 2), reported in humans, calves and other ruminants.
Other species and genotypes have been encountered, but much less frequently.
Humans and cattle are the most significant sources of Cryptosporidium. Sheep, pigs and horses
are also considered to be reservoirs (Olson et al., 1997). Rodents are not a significant source of
human-infective Cryptosporidium (Roach et al., 1993).
Oocysts are commonly found in water affected by human or livestock wastes by mechanisms
such as sewage, swimmer contamination and stormwater runoff. It is suggested that waterfowl
(ducks, geese) may be capable of picking up oocysts from their habitat and depositing them
elsewhere through discharge in their faeces. Typical concentrations in wastewater are on the
order of 1000–10 000 oocysts/L, whereas surface water concentrations in general range from
< 1 to 5000 oocysts/100 L (Guy et al., 2003).
The precise means through which Cryptosporidium causes human illness is not fully understood.
Damage caused by infection of red blood cells in the mucosa of the small intestine is known to
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contribute to illness. Cryptosporidium infection can result in illness of varying severity, ranging
from asymptomatic carriage to severe, life-threatening illness in immunocompromised
individuals. The primary characteristic of illness is profuse, watery and sometimes mucoid
diarrhoea. Other symptoms include nausea, vomiting, abdominal pain, weight loss, anorexia
and low-grade fever.
For Cryptosporidium, a variety of median infective doses have been reported—although, as is
the case with other pathogens, a single organism is theoretically sufficient to initiate infection.
Most (volunteer) feeding studies suggest that the median infective dose of Cryptosporidium is
between 80 and 140 oocysts (DuPont et al., 1995; Chappell et al., 1999, 2006; Okhuysen et al.,
2002). The prepatent period (time between ingestion and the excretion of new cysts) is roughly
4–9 days. Most healthy individuals experience a complete recovery, with the disease resolving
itself in about 1–2 weeks. Oocysts may continue to be shed in faeces for a short period following
recovery. Currently, there is no effective treatment for cryptosporidiosis in adults. The use of the
antimicrobial drug nitazoxamide has been approved by the U.S. Food and Drug Administration
for treatment in children (Health Canada, 2012b).
Epidemiology
Reported outbreaks of giardiasis and cryptosporidiosis from natural recreational waters have
been infrequent. Surveillance data published by the U.S. CDC for the period 1992–2002
indicated that Giardia was responsible for 9% (6 of 64) of the total number of outbreaks of
gastroenteritis reported for natural waters (Moore et al., 1993; Kramer et al., 1996; Levy et al.,
1998; Barwick et al., 2000; Lee et al., 2002; Yoder et al., 2004). Locations included recreational
lakes, a recreational river and a pond setting. Although Giardia has not been linked to outbreaks
in natural recreational waters in Canada, it is likely that cases have occurred that were not
detected or went unreported.
Surveillance data over the same period indicated that 6 (9%) of the 64 outbreaks of
gastrointestinal illness reported in natural waters were caused by Cryptosporidium (Moore et al.,
1993; Kramer et al., 1996; Levy et al., 1998; Barwick et al., 2000; Lee et al., 2002; Yoder et al.,
2004). Recreational lakes were recorded as the setting for the majority of the outbreaks. A large
outbreak at a New Jersey lake in 1994 involving 418 cases was the first recorded U.S. outbreak
of cryptosporidiosis related to recreational water use (Kramer et al., 1996). Treated recreational
water venues such as water parks and community and motel swimming pools have provided the
setting for the majority of outbreaks of cryptosporidiosis. Surveillance in Canada has been
limited; to date, however, there have been no reported outbreaks of cryptosporidiosis associated
with natural recreational waters. As with Giardia, it is expected that cases have occurred and
gone undetected or were not reported.
Relationship with indicators
Although the faecal indicators E. coli and enterococci are good indicators for enteric bacterial
pathogens commonly found in natural recreational waters, they have proven to be less effective
indicators of protozoan presence. Studies have demonstrated a lack of correlation between
concentrations of E. coli and enterococci and the presence of Giardia and Cryptosporidium in
surface waters (Hörman et al., 2004; Dorner et al., 2007; Sunderland et al., 2007). E. coli and
enterococci are present consistently in human and animal faeces, whereas the presence of
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Cryptosporidium and Giardia is source dependent. Additionally, shedding of these organisms
in the feces of recognized sources can be intermittent and seasonal. Giardia cysts and
Cryptosporidium oocysts are also more resistant to stresses than bacteria and may persist in the
environment for longer periods.
Other enteric protozoa of potential concern
Other enteric pathogenic protozoa such as Entamoeba and Toxoplasma can be shed in human
and animal faeces and thus can conceivably be present to contaminate recreational waters.
Currently, there have been no reported outbreaks involving these organisms in recreational
waters. Recreational water activity is not considered to be a significant risk factor for illness
caused by these organisms.
5.3.2 Free-living protozoa
The free-living protozoa recognized as the most important in natural recreational waters are
Naegleria and Acanthamoeba.
Naegleria fowleri
Naegleria are small, thermophilic, free-living freshwater amoebae. The genus Naegleria is
composed of six species. N. fowleri is the primary human pathogen and the species of concern
for recreational waters. The organism has a multi-stage life cycle consisting of a motile feeding
trophozoite stage, a non-replicating flagellate stage and an environmentally resistant cyst stage.
N. fowleri can be found worldwide in fresh water and soil. The organism has been isolated from
both natural and artificial water supplies, including lakes, rivers, hot springs, swimming pools,
hydrotherapy baths and tap water. No human or animal reservoirs have been identified. The
organism prefers warmer waters and can tolerate temperatures of 40–45°C (Percival et al., 2004).
Tropical and subtropical fresh waters and hot springs are particularly well suited for the survival
of N. fowleri. In colder waters, it is thought that the cyst form may survive in river and lake
sediments (Pond, 2005).
N. fowleri causes a disease of the central nervous system called primary amoebic meningitis
(PAM), which is almost always fatal. Human infection occurs when water containing the amoeba
is forcefully inhaled or splashed into the nasal passages (e.g., during diving, jumping, falling or
swimming underwater). Following inhalation, the organism travels through the nasal passages to
the brain, causing damage to the cells of the olfactory system and cerebral cortex. The onset of
illness is rapid. Symptoms include severe headache, high fever, intracranial pressure, stiff neck,
altered mental status and coma, ultimately leading to death. Treatment is possible but requires
prompt diagnosis and aggressive antimicrobial therapy. The organism is reported to exhibit
sensitivity to amphotericin B.
Naegleria have also been proposed as natural hosts for the bacterial pathogen Legionella.
Harbouring within Naegleria is thought to provide Legionella with an environment suitable
for replication, in addition to providing protection from environmental stresses.
Cases of PAM are extremely rare; it is estimated that in the United States, one case occurs for
approximately every 2.5 million swimmers (Visvesvara and Moura, 2006). Surveillance data
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indicated that 29 cases of the disease were reported in the United States over the period 1992–
2002, with an average occurrence of 0–6 cases per year (Moore et al., 1993; Kramer et al., 1996;
Levy et al., 1998; Barwick et al., 2000; Lee et al., 2002; Yoder et al., 2004). Reported outbreaks
have been limited to the southern United States, including Florida, Texas, Oklahoma, California,
Georgia and North Carolina. To date, there have been no recorded cases of PAM as a result of
recreational water contact in Canadian waters. The bulk of the evidence suggests that amoebic
meningoencephalitis is an unlikely health concern in Canada. However, researchers have
suggested that increasing lake temperatures brought on by climate change could result in an
expanded prevalence of this organism (Rose et al., 2001; Schuster and Visvesvara, 2004). Thus
the potential exists for this organism and disease to become an emerging concern for recreational
waters in the northern United States and Canada in the future.
Acanthamoeba
Acanthamoeba are small, free-living amoebae. The organisms possess a two-stage life cycle
consisting of a trophozoite feeding stage and an environmentally resistant cyst stage. The genus
Acanthamoeba contains approximately 20 species. A. culbertsoni, A. polyphaga and A.
castellanii are the species most commonly associated with human infections.
Acanthamoeba are considered ubiquitous in the environment. They can be found in fresh,
estuarine and marine waters, hot springs, soils and sewage, and human-made water supplies,
such as tap water and air conditioning condensers.
Pathogenic species of Acanthamoeba are responsible for two distinct clinical illnesses: amoebic
keratitis (AK), a painful, vision-threatening disease of the cornea caused by A. polyphaga and
A. castellanii; and granulomatous amoebic encephalitis (GAE), a fatal disease of the central
nervous system caused by A. culbertsoni. Infections occur via inhalation or through direct
contact with the mucous membranes of the eye or through abraded or traumatized skin.
Although pathogenic Acanthamoeba species can have a waterborne transmission, recreational
water contact is not considered to be a significant risk factor for either of these illnesses. The
principal risk for AK is poor hygienic practices in contact lens wearers (use of contaminated
solutions, inadequate disinfection practices). Infection can be acquired by wearing contact lenses
while swimming in lakes or ponds; however, the risk is considered extremely low. Currently,
recreational water contact is not considered a route for acquiring GAE.
Acanthamoeba have also been proposed as natural hosts for certain free-living bacterial
pathogens, namely Legionella and Mycobacterium. Survival within Acanthamoeba is thought to
provide these organisms with an environment suitable for replication, as well as to provide
protection from environmental stresses.
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6.0
Cyanobacteria and their toxins
Guideline values
The recommended guideline values for cyanobacteria and their toxins in recreational waters are:
Total cyanobacteria:
100 000 cells/mL
or
Total microcystins:
20 µg/L (expressed as microcystin-LR).
Exceedance of these values, or the development of a bloom indicates the potential for exposure
to cyanobacterial cells and/or their toxins in amounts which may, in some cases, be sufficient to
be harmful to human health. In general, contact with waters where a bloom exists or has very
recently collapsed should be avoided.
An appropriate monitoring program is advantageous to reduce the potential risk of user exposure
to cyanobacterial blooms and their toxins. It is advised that managed recreational water areas that
are suspected or are known to be susceptible to blooms be routinely monitored during the
bathing season. Authorities should visually monitor such supplies for cyanobacterial growth. A
swimming advisory may be issued at the discretion of the responsible authority.
In the event of bloom development, in order to fully characterize the extent of the risk posed by
the cyanobacterial population, it is further advised that authorities conduct sampling during, and
after the collapse of, the bloom. Should either of the guideline values be exceeded, a swimming
advisory may be issued by the responsible authority. When measuring the toxins it is important
that one measures total microcystins. “Total microcystins” includes both the microcystin that is
occurring free in the water and the microcystin that is bound to or inside the cyanobacterial cells.
Contact with waters where an advisory has been issued should be avoided until the advisory has
been rescinded. Published texts are available that can provide further information with respect to
the design and implementation of recreational water monitoring programs (e.g., Chorus and
Bartram, 1999)
Cyanobacteria
Cyanobacteria are bacteria that share features of both bacteria and algae. They are similar to
algae in size, possess blue-green pigmentation and are capable of photosynthesis; thus, they are
often termed blue-green algae (WHO, 2003a). Most planktonic cyanobacteria, including the
species found in Canadian lakes, form colonies, which can appear as irregular groupings of cells
or as filamentous chains that can be straight, coiled or branched (Chorus and Bartram, 1999;
Falconer, 2005). In a typical summer, a lake water sample can contain several species of
cyanobacteria, along with numerous other species of algae. Cyanobacterial cells contain small
gas bubbles called vacuoles, which allow them to control their buoyancy. The cells use this
buoyancy control to move up in the water column to where light is the greatest and down in the
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water column to where nutrients are more abundant (Falconer, 2005). In still, stratified surface
waters, cyanobacteria effectively use the light and nutrients to proliferate intensively, creating a
visible discoloration known as cyanobacterial blooms (Chorus and Bartram, 1999; Falconer,
2005). These blooms can be very dense and can have the appearance of being gelatinous or
resemble a collection of fine grass clippings or appear as a homogeneous, soupy mass, as if green
paint has been spilled into the water (WHO, 2003a; Falconer, 2005). Surface blooms or scums
can occur when the cells develop excess buoyancy and the water is calm enough to let them float
to the surface. This excess buoyancy develops when turbulence (e.g., during storms) sends the
cellsa too deep, during the hours of darkness, under carbon dioxide-limiting conditions, when the
population is at the end of the growth cycle or any combination of these factors. Offshore winds
may then drive these scums towards the shore where they can accumulate (Chorus and Bartram,
1999; Falconer, 2005). In this manner, cyanobacterial blooms may increase their density by a
factor of 1000 or more in a very short period of time (Chorus et al., 2000).
Cyanobacteria possess a number of special properties that determine their relative importance in
phytoplankton communities. Cyanobacteria have lower light intensity requirements and
demonstrate a greater affinity for nitrogen and phosphorus than do other algae and phytoplankton
and, thus, can outcompete these organisms under conditions where these factors might be
limiting (e.g., in turbid waters). The maximum growth rate for cyanobacteria has been reported
to occur at temperatures above 25°C (Chorus and Bartram, 1999). The combination of these
factors may help explain why blooms typically occur in the late summer months when waters are
warmer and the hours of daylight are beginning to grow shorter (Chorus and Bartram, 1999,
Falconer, 2005). As well, toxic cyanobacteria may not be grazed by zooplankton to the same
extent as other species of algae (Chorus and Bartram, 1999). More recently, it has been
suggested that certain mussel species (zebra mussels Dreissena polymorpha, quagga mussels D.
bugensis) selectively reject certain strains of Microcystis, possibly leaving toxic species to
proliferate under conditions of reduced competition (Brittain et al., 2000; Vanderploeg et al.,
2001).
The nutrient enrichment (eutrophication) of surface waters can also have a significant impact on
the frequency and severity of cyanobacterial blooms (Chorus and Bartram, 1999, Falconer,
2005). Nitrogen and phosphorus enter the environment as a result of both natural processes and
human activities (Chambers et al., 2001). Sources include runoff or erosion from naturally or
artificially fertilized soils, storm sewer discharges and discharges of agricultural, industrial or
sewage wastes (Jones and Armstrong, 2001). On-site wastewater disposal systems from isolated
dwellings (both urban and rural) can also be a significant source of nutrients.
Cyanobacterial toxins
Cyanobacterial blooms are considered a public health concern because direct contact with a
bloom can cause allergenic-like reactions; and some cyanobacterial species can produce toxins
that may have harmful effects on humans (Chorus and Bartram, 1999). Over 46 species of
cyanobacteria are capable of producing toxins (Sivonen and Jones, 1999). The most common
toxin-producing genera in fresh water are Anabaena, Aphanizomenon, Cylindrospermopsis,
Microcystis, Nodularia and Planktothrix (syn. Oscillatoria) (Falconer, 2005). Although the
conditions leading to the development of a bloom are relatively well known, the factors
responsible for the dominance of toxin-producing strains are not completely understood (Chorus
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and Bartram, 1999; Falconer, 2005). Consequently, toxin formation from cyanobacteria is even
less predictable than the cyanobacterial blooms themselves. Lakes that have never had a problem
can suddenly develop blooms that may contain toxins. Conversely, lakes that have shown toxic
blooms in the past may never show it again. It has been suggested that worldwide, an average of
60%of the cyanobacterial bloom samples investigated have been positive for cyanobacterial
toxins (range, 10–90%) (Chorus et al., 2000; WHO, 2003a). As a result, any bloom encountered
should be treated as potentially toxic. Cyanobacterial toxins for the most part are associated with
the cyanobacterial cells—either bound within membranes or occurring freely within the cells.
Toxin release to the surrounding waters can occur as the cells die or are damaged and leak their
contents (Chorus and Bartram, 1999). The bulk of the toxicity, if present, generally lasts as long
as the bloom; however, some toxin may still persist for a short period after the bloom is gone
(Chorus and Bartram, 1999; Falconer, 2005). Subsequently, contact with waters in which a
bloom has developed should be avoided until it can be unequivocally determined that there is not
a risk of contact with cyanobacterial toxins. The time of toxin persistence can depend on factors
such as the concentration before the bloom’s disappearance, and the efficiency of degradation by
natural microbial populations in the water (Falconer, 2005).
Some toxic benthic species (e.g., Lyngbya spp.) can grow to form dense, bottom-covering mats
of cyanobacterial material (Chorus and Bartram, 1999; WHO 2003a). These mats typically occur
in clear, shallow waters where sunlight can penetrate to the bottom. The mats can occasionally
be dislodged and washed ashore, where they may be scavenged by animals. Human health risks
from these mats are considerably less than those from scums produced by other cyanobacterial
species; however, they can still present a risk for domestic pets and livestock (WHO 2003a).
There are other known toxic marine algal species that are capable of forming blooms (e.g.,
Alexandrium spp. and the phenomenon known as “red tide”) and producing toxins (i.e., those
toxins responsible for shellfish poisonings) (Chorus and Bartram, 1999). However, as the focus
of this section is the human health risks from exposure to toxic cyanobacteria through
recreational water activities, they will not be discussed here.
Confirmation of toxins within a bloom cannot be accomplished by visual inspection; samples
must be sent to a laboratory for analysis. Warning signs may be observed, such as the presence
of dead waterfowl or other wildlife along the shoreline or reports of domestic animal poisonings
(specifically cattle and dogs) (Chorus and Bartram, 1999). Still, toxic blooms can occur without
any noticeable effect on the local animal populations.
Increased awareness of cyanobacterial blooms, coupled with occasional escalations in bloom
detection, has prompted a growing concern over the possible development of blooms in
recreational waters and the resulting health implications for recreational water users. Both
drinking water and recreational water are potential sources of exposure. In general, because of
circumstances such as the seasonality and localized nature of blooms, their unappealing aesthetic
properties and the way drinking water supplies and monitored recreational water areas are
managed, the likelihood of exposure to cyanobacterial toxins in sufficient amounts to constitute a
chronic or acute health risk is considered to be relatively low. Under circumstances where a
recreational area is experiencing prolonged and persistent blooms and where intensive
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recreational activities are continuing, the risks of acute exposure may be greater (Funari and
Testai, 2008).
Direct epidemiological evidence of adverse health effects associated with recreational exposure
to cyanobacteria is limited. Information on human symptoms of illness have come largely from
anecdotal and case reports, many of which have gaps in the information on cyanobacterial
conditions at the time of exposure (e.g., species present, cell and/or toxin concentrations). The
symptoms most frequently described in these reports have been hay fever-like symptoms,
gastrointestinal symptoms and skin irritations (Stewart et al., 2006a). Additional information has
been provided from toxicological studies that used animal models, as well as from occasional
accounts of animal poisonings.
There are three main routes for human exposure to cyanobacteria and their toxins in recreational
waters: ingestion, inhalation and direct body contact (Chorus and Bartram, 1999). Cases of
illness have been reported that have provided anecdotal evidence of toxicity to recreational water
users in bloom-impaired waters through the accidental swallowing of water (Chorus and
Bartram, 1999; Stewart et al., 2006a). There has also been experimental evidence to suggest that
inhalation of contaminated aerosols may be equally important as a potential route of exposure
(Fitzgeorge et al., 1994; Chorus and Bartram, 1999). This route would be relevant for activities
in areas where aerosols are generated, such as waterskiing. Activities involving sudden or
repeated immersion of the head (such as diving, windsurfing or kayaking) may also lead to
ingestion or inhalation exposure via water forced into the mouth and/or nasal passages. Direct
contact with cyanobacterial populations has also been known to cause irritative effects of varying
severity, although the exact mechanisms for this are not fully understood. Some allergic
reactions have been reported in sensitive individuals. It has been suggested that the irritations are
due to unknown cyanobacterial components, separate from the toxins (Chorus and Bartram,
1999). Bathing suits and wet suits may also function to exacerbate the potential for skin
irritations by trapping the cells and then disrupting their contents as a result of the friction
created between the suit material and the user’s skin (Chorus and Bartram, 1999).
There are several known cyanobacterial toxins that can pose concerns for recreational water
users. These include microcystins, nodularins, anatoxins, cylindrospermopsin, dermatotoxins
and irritant toxins (Chorus and Bartram, 1999). Microcystins and nodularins are cyclic peptides
that affect the liver (hepatotoxins), anatoxins are alkaloids that target the nervous system
(neurotoxins) and cylindrospermopsin is an alkaloid that affects a wide range of organs (general
cytotoxin) (Chorus and Bartram, 1999). The dermatotoxins (alkaloids) and irritant toxins
(lipopolysaccharides) are toxins that cause irritations of exposed tissues (Chorus and Bartram,
1999).
Microcystins
Microcystins are hepatotoxins that disrupt the functioning of enzymes called protein
phosphatases, which are important metabolic switches in human and animal cells (WHO, 2003a).
Their primary target is the liver, with the main route of entry into cells occurring through a
membrane transport mechanism known as the bile acid carrier (Chorus and Bartram, 1999).
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Microcystins are produced by most species of Microcystis and some species of Anabaena—two
notable scum producers (WHO, 2003a). Other cyanobacteria capable of producing the toxin are
Oscillatoria (syn. Planktothrix), Nostoc and Anabaenopsis. Over 70 microcystin variants have
been isolated from bloom samples (Sivonen et al., 1992). Variants are named according to the
variable amino acid position encountered in their structure. Microcystin-LR, the most commonly
encountered variant, is named for possessing the amino acids leucine (L) and arginine (R) in the
variable position (Chorus and Bartram, 1999). Microcystins and, more specifically, microcystinLR represent the most frequently encountered cyanotoxins in cyanobacterial blooms in temperate
surface waters in North America, based on available monitoring data. Therefore, they are the
cyanotoxins of most relevance for recreational waters in Canada.
Symptoms reported from incidences of human recreational water exposures to waters
contaminated by blooms of Microcystis and Anabaena have included headaches, nausea,
vomiting, diarrhoea, abdominal pain, muscle aches, fever, mouth ulcers, blistering of the lips,
sore throat, skin rashes and ear and eye irritations (Chorus and Bartram, 1999). Cases of animal
poisonings (cattle, sheep and canines) resulting from contact with Microcystis blooms have
provided evidence of liver toxicity and have included fatalities, substantiating the concern for
human health effects from exposure. To date, no human fatalities have been reported as a result
of exposure to microcystins through recreational water activities.
Toxicological studies using animal models have been used to provide further evidence of
possible human health effects (Chorus et al., 2000). Microcystin-LR has been shown to be toxic
following acute exposures in rodents. The oral (by gavage) LD50 is 5000 μg/kg body weight (bw)
in mice and higher in rats (Fawell et al., 1999). Studies of acute and short-term exposures have
demonstrated evidence of liver enzyme changes and tissue damage (inflammation,
haemorrhaging, lesions) in mice as a result of exposure to microcystin-LR through oral
administration (Fawell et al., 1999). Chronic exposures to low levels of microcystins have been
shown to lead to progressive liver injury in experimental studies in pigs and mice (Chorus and
Bartram, 1999). Microcystins are thought to be capable of promoting tumour development by
interfering with the normal mechanisms of cell division. Evidence has been provided on the
ability of microcystin-LR to promote the growth of certain types of tumours in mice subjected to
prolonged exposure to the toxin through oral administration (Falconer, 2005). Presently there is
insufficient evidence for the carcinogenicity or genotoxicity of microcystin-LR (Funari and
Testai, 2008; Gaudin et al., 2009). However, the toxin has been listed as “possibly carcinogenic
to humans” by the International Agency for Research on Cancer as a result of its tumourpromoting potential (IARC, 2010).
Both toxic and non-toxic species exist for all of the predominant microcystin-producing genera,
and it is thought that different species or strains can vary in their toxic potential (Chorus and
Bartram, 1999; Carillo et al., 2003; Dittmann and Börner, 2005). As a result a single bloom may
consist of a mixture of non-toxic and variably toxic strains. It is generally felt that variations in
toxicity within a bloom are due to the rise and fall of subpopulations of strains having different
toxic potential. Environmental factors (e.g., conditions of winds, sunlight and temperature) are
also thought to contribute to this phenomenon, but less significantly (Chorus and Bartram, 2003).
Recently, the genes responsible for microcystin production (mcy genes) have been identified and
sequenced (Falconer, 2005). Numerous studies have subsequently confirmed the detection of the
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mcy gene cluster as a tool to discriminate between toxic and non-toxic strains of Microcystis,
Anabaena and Planktothrix in both laboratory samples and field isolates (Dittmann and Börner,
2005).
Nodularins
Nodularins are hepatotoxins found in blooms caused by strains of the brackish-water species
Nodularia spumigenia. The toxins are closely related to microcystins in both structure and
function and thus act similarly by inhibiting protein phosphatase activity in liver cells (Chorus
and Bartram, 1999).
Data derived from experimental studies, although limited, have suggested that nodularin exhibits
toxicity similar to that of microcystin-LR. Results obtained from chronic toxicity studies using
animal models have suggested that nodularins may be a more potent tumour promoter than the
microcystins (Chorus and Bartram, 1999).
Nodularia blooms have been encountered in brackish lakes in Australia and New Zealand, as
well as in the Baltic Sea. In general, blooms in fresh water are considered extremely rare (Chorus
and Bartram, 1999). To date, there have been no recorded occurrences of Nodularia blooms in
North American waters. As a result, they are not considered to be a significant public health
threat in Canadian recreational waters.
Anatoxins
The anatoxins (anatoxin-a, anatoxin-a(S), homoanatoxin-a) are neurotoxins that can be found
in blooms produced by Anabaena (anatoxin-a, anatoxin-a(S)), Oscillatoria (anatoxin-a,
homoanatoxin-a) and Aphanizomenon (anatoxin-a). Anatoxins interfere with the activity of
the nerve transmitter acetylcholine, which affects the functioning of the nervous system by
disrupting communication between neurons and muscle cells (Chorus and Bartram, 1999). Acute
toxicity is characterized by paralysis of both the skeletal and respiratory muscles, resulting in
tremors, convulsions and, ultimately, death due to respiratory failure (Rogers et al., 2005).
Information on health effects associated with anatoxins has been gleaned from occasional
accounts of animal poisonings and toxicological investigations. Anatoxin-a has been the most
widely studied member of this group. Limited data are available for homoanatoxin-a and
anatoxin-a(S) (Chorus and Bartram, 1999). However, it is generally regarded that although the
toxins have somewhat different mechanisms of action, each is capable of causing fatalities at
elevated doses. The oral LD50 of anatoxin-a is reported to be greater than 5000 µg/kg bw
(Fitzgeorge et al., 1994). Anatoxin-a was shown to be toxic in experimental mice upon acute
exposure to high levels of the toxin (15 000 µg/kg bw). Death was reported to occur within
minutes (Chorus and Bartram, 1999). Conversely, in a number of experimental studies, recovery
after exposure to sublethal doses of anatoxin-a was complete, with no signs of clinical toxicity.
To date, there has been only one reported fatality resulting from exposure to cyanobacterial
neurotoxins in natural waters. Exposure occurred through ingestion of contaminated water during
accidental immersion at a location where swimming was not permitted (Falconer, 2005).
Blooms of anatoxin-producing species are not routinely reported in North American waters
and are considered to occur far less frequently than those of the microcystin-producing
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cyanobacteria. In addition, the anatoxins are relatively unstable and, as such, are not considered
to be as widespread as microcystins in water supplies (Chorus and Bartram, 1999). As a result,
anatoxins are currently considered to be of lesser concern than microcystins for Canadian
recreational waters.
Cylindrospermopsin
Cylindrospermopsin is a relatively new cyanobacterial toxin named after the species from which
it was first isolated: Cylindrospermopsis raciborskii. It is a member of the alkaloid family of
toxins; however, the mode of action is considerably different from that of the anatoxins.
Similarly, cylindrospermopsin does exhibit hepatotoxic activity, but operates by a mechanism
much different from that of the other microcystins (Chorus and Bartram, 1999). The toxin has
been isolated from certain species of a few other genera of cyanobacteria, including Anabaena
and Aphanizomenon (Falconer and Humpage, 2006)
Cylindrospermopsin is considered to be a general cytotoxin, which acts by inhibiting protein
synthesis. Cellular damage is caused by blocking the functioning of key proteins and enzymes.
The liver and kidneys are considered the main targets for the toxin; however, crude extracts of
Cylindrospermopsis given to mice have also demonstrated evidence of injury to other organs,
such as the lung, spleen, thymus and heart (Chorus and Bartram, 1999). Data, derived from a
study in which mice orally dosed with cylindrospermopsin demonstrated a positive but not
statistically significant increase in the number of tumours, have provided evidence that the toxin
may also possess carcinogenic properties (Falconer and Humpage, 2001).
The first recorded incident of human cylindrospermopsin poisoning occurred in 1979. Initially
termed “Palm Island Mystery Disease,” this outbreak of hepatoenteritis among residents of a
tropical island off the coast of Queensland, Australia, was later attributed to a bloom of
C. raciborskii in the drinking water reservoir. Patients had reported symptoms that included
vomiting, malaise, headache and constipation, later followed by bloody diarrhoea. Blood and
urine analysis revealed evidence of liver and kidney damage; however, all patients were reported
to have recovered following treatment. Follow-up research led to the identification of the toxin
cylindrospermopsin produced by this cyanobacterial species. At present, there have been no
human fatalities associated with cylindrospermopsin, and there have been no other recorded
poisonings since the Palm Island outbreak (Chorus and Bartram, 1999).
Cylindrospermopsis is more frequently encountered in the warmer waters found in tropical and
subtropical locations of the world. Blooms have been routinely encountered in freshwater lakes
and drinking water reservoirs in Australia, South and Central America, and the state of Florida.
To date, the organism has been only occasionally encountered in temperate fresh waters. In
North America, populations have been detected in several northern U.S. states (Michigan, Ohio,
Minnesota, Illinois and Indiana), as well as in the province of Manitoba.
There are several notable differences between populations of Cylindrospermopsis and those of
cyanobacteria having microcystin-producing potential such as Microcystis or Anabaena.
Cylindrospermopsis does not form slicks or scums of the type produced by these organisms. In
Cylindrospermopsis blooms, the regions of highest cell concentrations are located beneath the
surface (Falconer, 2005). As well, in these blooms, a significant portion of the toxin is released
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into the surrounding water, in contrast to the situation in microcystin-producing species in which
the toxin is predominantly retained within the cells and released only upon cell rupture or death
(Falconer, 2005). Blooms of Cylindrospermopsis are considered infrequent in Canadian waters.
Those caused by microcystin-producing species are far more prevalent. Nonetheless, the
increasing frequency at which this organism is being detected in temperate fresh waters has
resulted in the identification of Cylindrospermopsis blooms as a potentially emerging concern for
recreational waters in Canada and the United States.
Saxitoxins
Saxitoxins are neurotoxins that belong to a larger family of toxins referred to as Paralytic
Shellfish Poisoning (PSP) toxins. These toxins, originally isolated from shellfish having fed on
toxic species of marine dinoflagellates, have also been isolated from several genera of freshwater
cyanobacteria, including Aphanizomenon, Anabaena, Cylindrospermopsis, Lyngbya and
Planktothrix (Oscillatoria) (Chorus and Bartram, 1999; Aráoz et al., 2010). Saxitoxins act by
blocking ion channels in nerve and muscle cells, which prevents the transmission of electrical
impulses. This can lead to neuromuscular paralysis, which can ultimately be fatal owing to the
resulting respiratory failure.
Most of the toxicological information on saxitoxins has been obtained from studies using toxins
produced by the marine dinoflagellates (e.g., Alexandrium spp.). Nevertheless, saxitoxin types
derived from different sources remain identical in structure and toxicological profile (Funari and
Testai, 2008). Saxitoxin is considered one of the most toxic of the PSP toxins (Chorus and
Bartram, 1999). The oral LD50 reported for saxitoxin is 263 µg/kg bw (Mons et al., 1998; Funari
and Testai, 2008). Animal deaths have been linked to contact with cyanobacterial blooms
containing saxitoxin (Negri et al., 1995). There have been no saxiotoxin-related illnesses
reported for humans through drinking water or recreational water exposure (Chorus and Bartram,
1999; Aráoz et al., 2009).
Saxitoxin-containing blooms are considered widespread in Australia, and toxic blooms have also
been detected in waters in Brazil and in the southern and northern United States (Chorus and
Bartram, 1999; dos Anjos et al., 2006). Presently, saxitoxins are not considered to be as
significant a concern as microcystins in Canadian recreational waters. However, the detection of
saxitoxins in blooms in temperate fresh waters suggests that this issue should continue to be
monitored.
Dermatotoxins and other irritant toxins
Certain marine cyanobacteria such as Lyngbya, Oscillatoria and Schizothrix can produce toxins
called aplysiatoxins and lyngbyatoxins, which have been reported to cause severe dermatitis
among swimmers who come in contact with the cyanobacterial filaments. Aplysiatoxins are
considered potent tumour promoters and are also thought to demonstrate other properties that
may be linked to carcinogenesis (Chorus and Bartram, 1999). As these are primarily marine
species, they are not thought to be of concern for freshwater lakes and rivers.
Although not well understood, it is also thought that the lipopolysaccharide component of
the cyanobacterial cell wall can elicit an irritant or allergenic response in humans.
Lipopolysaccharides have been known to exhibit pyrogenic (fever-inducing) and toxic
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properties. It is generally regarded, though, that lipopolysaccharides from cyanobacteria are
considerably less toxic than those of other Gram-negative bacteria, such as Salmonella (Chorus
and Bartram, 1999). Nonetheless, it is thought that they may be at least partially responsible for
some of the non-specific irritative effects associated with human exposure to cyanobacterial
blooms.
Compound of interest: Β-methylamino-L-alanine (BMAA)
An emerging topic of interest involves the unusual amino acid Β-methylamino-L-alanine
(BMAA), its links to cyanobacteria and the research findings concerning its potential neurotoxic
capabilities. The present evidence does not suggest that BMAA is a recreational water quality
hazard of human health concern. Information is provided on the current state of evidence on this
issue.
BMAA was first identified during exploratory studies into the high rate of amyotrophic lateral
sclerosis/parkinsonism–dementia complex (ALS-PDC, a neurodegenerative disease with
symptoms similar to Parkinson’s disease and Alzheimer’s disease) observed among the
Chamorro people of Guam. Findings reported by the researchers during investigations into this
issue included that BMAA could be detected in brain tissues of ALS-PDC patients in Guam and
Canada (Cox et al., 2003), and that BMAA could be found in virtually all groups of
cyanobacteria including the notable freshwater genera: Anabaena, Aphanizomenon, Microcystis,
Nodularia and Oscillatoria (Cox et al., 2005; Metcalf et al., 2008). The further observation that
BMAA could be detected in flying foxes (i.e., bats, which constitute a portion of the Chamorro
diet) led the researchers to hypothesize that BMAA could be subject to magnification up the food
chain.
These findings and the proposed implications are currently the subject of debate (CRC 2005).
Some researchers have provided conflicting evidence regarding BMAA detection in brain tissue
of ALS-PDC patients (Montine et al., 2005), while others have questioned the data relating to
neurotoxicity and food chain magnification (Duncan and Marini, 2006).
Research into this issue is considered preliminary, requiring further investigation. It is proposed
that much more work is needed before a cause and effect relationship between BMAA and
neurological disease can be established or discounted (CRC, 2005). Similarly, there is
insufficient evidence at this time to suggest that water supplies or dietary sources could
constitute a significant source of BMAA exposure. Developments on this topic will continue to
be monitored.
Guideline rationale
The recommended guideline values for total cyanobacteria and total microcystins in Canadian
recreational waters are based upon the approach used in the derivation of the maximum
acceptable concentration (MAC) for microcystin-LR in the Guidelines for Canadian Drinking
Water Quality (Health Canada, 2002). The use of single guideline values for total cyanobacterial
cell density and total microcystin concentration is the preferred approach to setting guidelines for
Canadian recreational waters at this time. They are intended to protect against both the risk of
exposure to microcystins through inadvertent ingestion of water as well as from other harmful
effects that may be possible after exposure to high densities of cyanobacterial material. In this
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situation, values were derived based on children as they may be more intensely affected because
they spend more time in water than adults and are more likely to accidentally swallow
contaminated water. The guideline value for total microcystins (expressed as microcystin-LR) is
intended to be protective against exposure to other microcystin variants that may be present.
For microcystin-LR, the tolerable daily intake (TDI) for recreational water exposures is derived
as follows:
TDI =
40 µg/kg bw per day
= 0.4 µg/kg bw per day,
100
where:
• 40 µg/kg bw per day is the NOAEL for liver changes derived from the 13-week mouse study
conducted by Fawell et al. (1999); and
• 100 is the uncertainty factor (×10 for intraspecies variation, ×10 for interspecies variation).
An additional uncertainty factor for less-than-lifetime study was not considered necessary, as the
types of exposures being considered are short term and episodic in nature.
The guideline value for total microcystins is calculated from the TDI for microcystin-LR as
follows:
0.4 µg/kg bw per day × 13 kg bw
Guideline value =
0.25 L/day
≈
20 µg/L for total microcystins
where:
• 0.4 µg/kg bw per day is the TDI, as derived above;
• 13 kg bw is the average weight of a child aged 7 months to 4 years (Health Canada, 1994);
and
• 0.25 L/day (250 mL/day) is the estimated amount of water accidentally ingested per day
during recreational water activities by a child.
The value used for the amount of water accidentally ingested per day is probably a conservative
estimate. Risk assessment models developed for enteric pathogens in recreational waters have
typically assumed a value of 100 mL for the amount of water likely to be swallowed through
recreational water activities (Haas, 1983; Gerba et al., 1996; Mena et al., 2003). Yet the basis for
this appears to be more historical than empirical. WHO (2003a) suggests that a child may be
capable of consuming 250 mL of water during extensive playing. During an empirical
investigation of the volume of water swallowed during swimming activities, Evans et al. (2006)
reported average values of 24 mL for adults (95% range, 2–84 mL) and 47 mL for children (95%
range, 3–142 mL). The authors further observed that some swimmers swallowed up to 280
mL/hour.
In addition to being capable of initiating toxic effects through ingestion, microcystins have been
suspected to be associated with other irritative effects that can occur through recreational water
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contact, such as injuries to the tissues of the mouth and lips. As a result, it is possible that a more
conservative estimate is appropriate as a worst-case scenario to account for the total amount of
water ingested, and not just the portion accidentally swallowed. The value of 250 mL represents
a risk management decision, derived based on the assessment of the available information
regarding the likely risk of ingestion and of the potential risks for the recreational water user.
The guideline value for cyanobacterial cell density (expressed as total cyanobacteria) represents
a general indication of the potential for bloom development and is intended to be protective
against exposure to high densities of cyanobacterial material. It similarly may be used to provide
protection against exposure to blooms of other cyanobacteria with toxic potential, and not just
the microcystin-producing species. It is calculated based on the microcystin guideline value
using the recognized reference value for the average toxin quota for Microcystis cells and reflects
the highest likely water quality hazard scenario of a toxic bloom containing high levels of
microcystin (WHO, 2003a; NHMRC, 2008):
20 µg/L × 10−3 L/mL
Guideline value =
=
2 × 10−7 µg/cell
100 000 cells/mL for total cyanobacteria
where:
• 20 µg/L is the guideline value for total microcystins in recreational waters, as derived in the
previous section;
• 2 × 10−7 µg/cell is the toxin cell quota for total microcystins per cell (WHO, 2003a;
NHMRC, 2008); and
• 10−3 L/mL is the factor converting litres to millilitres.
Currently, there is insufficient evidence to derive recreational water guidelines for other
cyanobacterial toxins, such as the anatoxins or cylindrospermopsin. Based on the assessment, it
was similarly determined that there is not sufficient data upon which to base a guideline whose
intent is to provide protection from the risk of allergenic or other irritative effects caused by
unknown cyanobacterial substances. Studies conducted by Pilotto et al. (1997, 2004) on the
effects of human skin contact with cyanobacteria were reviewed. Both studies reported that skin
contact with cyanobacteria across a wide range of cell densities resulted in reactions of only mild
severity in only a very small proportion of subjects. No dose–response relationships or thresholds
for irritation could be observed. It was judged that these findings did not provide sufficient
argument for the need for a separate guideline.
This approach and guideline value for microcystin are consistent with criteria derived by the
province of Québec for microcystin-LR toxic equivalents in recreational waters: 16 µg/L
(INSPQ, 2004). Further, the guideline value for total cyanobacteria is consistent with the WHO
guideline of 100 000 cyanobacterial cells/mL for the moderate probability of adverse health
effects. In its supporting documentation, WHO (2003a) further proposes that at this
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cyanobacterial density, a concentration of 20 µg microcystin/L is likely if the bloom consists of
Microcystis with an average toxin content of 2 × 10−7 µg/cell.
The Canadian approach is also similar in principle to the Australian Level 1 Guidelines
(NHMRC, 2008); however, a different animal model and derivation process have been used in
the Canadian calculations. The Australian approach was based on data considered to be the most
suitable for deriving guidelines for toxic cyanobacteria in Australian recreational waters.
Occurrence in the environment
In a 1995 survey of 16 recreational water bodies in southern Manitoba, Jones et al. (1998)
reported that microcystin-LR could be found at 44% of the sites, with concentrations occurring
in the 0.1–0.6-µg/L range. The authors further reported that cyanobacterial cell density was not a
good predictor of microcystin-LR and that there was no correlation between toxin concentration
and the other environmental variables monitored in the study.
Kotak et al. (1996) reported on a 3-year investigation that measured microcystin concentrations
encountered in phytoplankton and surface water samples collected from four freshwater lakes in
central Alberta. Phytoplankton concentrations of microcystin were found to be highest in the two
most hypertrophic lakes (Driedmeat and Little Beaver lakes). The highest microcystin
concentrations were observed in August and September, and were reported to correspond to
periods where Microcystis cell counts were at their highest (>200 000 cells/mL). Levels reported
during these months at one of the lakes (Driedmeat) ranged from 1.2 to 6.1 µg/L. The highest
microcystin concentration observed during the course of the study was 11 µg/L in a sample
collected from Little Beaver Lake in mid-August.
Giani et al. (2005) reported total microcystin concentrations of 0.008–1.91 µg/L (mean,
0.140 µg/L) in water samples collected during a survey of 22 freshwater lakes in southern
Québec. None of the lakes were reported to be affected by cyanobacterial blooms at the time of
sampling. The authors reported that all lakes contained detectable levels of cyanobacterial genera
having toxic potential (including Microcystis, Anabaena and Oscillatoria), although these were
typically encountered at concentrations considered far below levels of concern.
Rinta-Kanto et al. (2005) reported on the characterization of two separate bloom events that had
developed in the western basin of Lake Erie: the first in August 2003 and the second in August
2004. Satellite imaging was used by the researchers to detect the presence of the blooms and was
similarly applied to identify potential sites for analysis. Bloom samples were collected and
analysed for total microcystin concentrations (cell-bound fraction). At the same time, estimates
of the presence of total and toxic Microcystis concentrations were determined by QPCR analysis,
with targets specific for a Microcystis 16S rDNA gene fragment and the microcystin synthase
gene mcyD. Microcystin concentrations encountered in the 2003 bloom ranged from <0.3 to 15.4
µg/L, and estimates of the corresponding Microcystis concentration ranged from below the
quantifiable limits of detection to a peak value of 3.9 × 108 Microcystis equivalents/L. In the
bloom encountered the following year, microcystin concentrations ranged from 0.1 to 2.6 µg/L,
and estimates of the total Microcystis concentration spanned from approximately 5 × 103 to
3 × 106 Microcystis equivalents/L.
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Related epidemiology
No human cyanotoxin-related deaths have been associated with exposure via recreational water.
To date, there have been only two reported instances of human fatalities as a result of exposure
to cyanobacterial toxins. The most notable are those associated with the human dialysis tragedy
that occurred in Caruaru, Brazil, in 1996 (Jochimsen et al., 1998). In total, 56 patients at a
haemodialysis clinic died as a result of liver failure from exposure to microcystins in the water
used for dialysis. The presence of toxic cyanobacteria in the city’s drinking water reservoir and
insufficient water treatment were cited as major factors contributing to the deaths. More recently,
in 2002, a Wisconsin teenager died following exposure to a toxic scum in a pond on a public golf
course where swimming was not permitted (Falconer, 2005). Blood and stool samples collected
revealed the presence of both the cyanobacterial species Anabaena flos-aquae and the neurotoxin
anatoxin-a.
There have been few reported accounts of human illness as a result of contact with toxic
cyanobacterial populations in recreational waters worldwide. The predominant health effects
encountered from such exposures may be gastrointestinal or flu-like in nature and thus may often
go unreported or are attributed to other causes (Falconer, 2005). As well, most of the cases of
illness are investigated retrospectively; thus, information regarding the exact conditions of
exposure (number and type of organisms, identification and/or concentration of toxins) is rarely
available.
In Saskatchewan in 1959, despite warnings and reports of livestock deaths, people swam in a
lake infested with cyanobacteria. Thirteen individuals in total became ill, reporting symptoms of
nausea, muscular pain, headache and diarrhoea (Dillenberg and Dehnel, 1960). In England in
1989, 10 of 20 soldiers became ill after canoe training and swimming in water affected by a
Microcystis bloom. Swimming skills and the amount of water ingested were reported to have
been related to the degree of illness (Turner et al., 1990). In 1991, in Australia, two teenage girls
developed gastroenteritis and muscle pains after swimming in the Darling River near Wilcannia
during a cyanobacterial bloom containing Anabaena (NHMRC, 2008). In the same year, two
cases of conjunctivitis and one of dermal irritation were attributed to swimming during an
Anabaena circinalis bloom in Lake Cargelligo, Australia (NHMRC, 2008).
Pilotto et al. (1997) reported on the results of a prospective epidemiological study designed to
investigate the health effects encountered in swimmers after contact with cyanobacteria in
recreational water. Symptoms were shown to increase significantly with the duration of water
contact and increasing cyanobacterial cell density, but did not correlate with cyanotoxin
concentrations. Participants in this study who were exposed to 5000 cells/mL for more than 1
hour had a higher rate of symptom occurrence than did unexposed participants. However, after
controlling for previous exposure and prior illnesses, the data suggested that only a small number
of people were affected, and only with mild irritation. Stewart et al. (2006b) conducted a
prospective epidemiological study to examine the incidence of acute symptoms in recreational
water users exposed to different amounts of cyanobacteria based on measurements of surface
area per volume (SA/V). The authors observed that symptoms were more likely to be reported
among persons exposed to high (>12 mm2/mL) versus low (<2.4 mm2/mL) SA/V levels, and that
respiratory symptoms were recorded with the greatest frequency. Cyanotoxins were detected
only occasionally and at low SA/V levels during the study.
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Managing health risks
A multi-barrier approach is considered the best strategy to reduce the risk of exposure to
cyanobacteria or their toxins in recreational waters. This approach combines the use of the
recommended indicators of water quality alongside actions both to reduce the extent of the water
quality hazard and to restrict swimmer exposures during periods or in areas perceived to be of
increased risk.
The occurrence of cyanobacterial blooms in recreational waters is extremely difficult to predict.
Bloom development is influenced by a variety of physical, chemical and biological factors. As a
result of the interplay of these factors, there may be large year-to-year fluctuations in the levels
of cyanobacteria and their toxins (Health Canada, 2002). Managed recreational water areas that
are suspected or are known to be susceptible to blooms should be routinely monitored during the
bathing season. Authorities should visually monitor such supplies for cyanobacterial growth.
Toxicity within a bloom can vary considerably, particularly within large blooms. Further, a water
body in which a bloom has developed may still contain toxins for a short period after the bloom
has dissipated. In order to fully characterize the extent of the risk posed by the cyanobacterial
population, authorities should conduct sampling during, and after the collapse of, the bloom.
Collection of multiple samples may be required to identify any regional or localized differences
in cyanobacterial cell density and toxin concentration. It is advised that both total cyanobacterial
cell densities and total microcystin concentrations be monitored as part of a risk management
strategy for cyanobacteria and their toxins in Canadian recreational waters. As discussed,
monitoring of microcystin levels is necessary to determine the potential health risk posed by a
cyanobacterial population, while cell count determinations are useful for providing a general
indication of the potential for bloom development and, thus, the possible presence of
cyanobacterial toxins.
Waters shown to exceed the established guideline values or those in which a bloom has
developed may result in human exposure to cyanobacterial material or cyanotoxins in amounts
sufficient to be harmful to human health. A swimming advisory may be issued at the discretion
of the responsible authority. Contact with waters where an advisory has been issued should be
avoided until the advisory has been rescinded. Further information on the posting of recreational
waters can be found in Part I (Management of Recreational Waters).
Other barriers for risk reduction can include the provision of educational materials outlining
steps the public may take to reduce their personal risk in the event of a bloom. Guidance
provided in materials for public communication may include the following recommendations:
•
Recreational water users should avoid areas with visible cyanobacterial blooms and/or
scums, as all blooms have the potential to be toxic. Direct contact with bloom material or
accidental ingestion of contaminated waters can be harmful to recreational water users.
Inhalation may also be an important route of exposure to cyanobacterial toxins during
activities in bloom-affected areas where aerosols might be generated.
•
Users should shower or wash themselves as well as any item that may have accidentally
come into contact with cyanobacterial material as soon as is practical upon exiting the water.
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•
Any user experiencing adverse health effects from recreational water activity should consult
a medical professional and, if necessary, alert the appropriate authorities.
•
Users should not let pets swim in or drink from areas where the water has taken on an
abnormal discolouration consistent with that of a bloom, or where accumulations of
cyanobacterial material are visible. Should pets come in contact with bloom-affected waters,
they should be rinsed off immediately to remove all traces of cyanobacterial material that
may accidentally be ingested.
The use of algicides is not recommended as a measure to control cyanobacterial populations. The
addition of copper sulphate or other algicides to mature toxic blooms may have the effect of
destroying the cyanobacterial cells; however, this action may also cause the release of significant
amounts of cyanotoxin into the surrounding waters if present within the cells. Jones and Orr
(1994) reported that microcystin-LR could be detected up to 21 days after algicide treatment of a
toxic Microcystis aeruginosa bloom that had developed in a recreational lake. Environmental
concerns have also been cited as additional reasons for not pursuing this approach, as the
algicides can be detrimental to the healthy functioning of the aquatic ecosystem.
Longer-term risk management actions that may also be considered as barriers to reduce the
impact of toxic cyanobacteria in surface waters may involve the identification of the major
nutrient inputs, as well as the development of strategies to reduce nitrogen and phosphorus
loading through effective control of agricultural, municipal sewage and residential waste disposal
practices. Tracking total phosphorus levels in surface waters has been suggested as a proactive
step for recognizing water bodies that have the potential for bloom development (Chorus and
Bartram, 1999).
Summary
1. The development of cyanobacterial blooms in waters that are used for recreational purposes
is dependent on many factors that can be difficult to predict. Blooms can develop very
rapidly under the appropriate conditions, and lakes that have never had a problem can
suddenly become toxic.
2. Serious swimmer illnesses have been reported following exposure to toxic cyanobacterial
blooms in recreational waters. A water body containing a cyanobacterial bloom may still
contain toxins after visible signs of the bloom have disappeared. In general, contact with
waters where a bloom exists or has very recently collapsed should be avoided. Contact with
waters where an advisory has been issued should be avoided until the advisory has been
rescinded.
3. To date, there have been very few reported cases of cyanobacterial toxin-related illness due
to recreational water activity in Canadian or international waters. The combination of visual
inspection, water quality monitoring and public notification and education alongside actions
and procedures for reducing nutrient inputs represents the most effective approach to
protecting the health of recreational water users.
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7.0
Other biological hazards
Guideline values
No guideline values can be specified for any of the organisms identified in this section. In
general, areas used for recreational water purposes should remain as free as is practical from
these organisms. Furthermore, recreational water activities should not be pursued in waters
where the responsible authority deems their presence poses a risk to the health and safety of the
users.
This section provides guidance on other organisms that may affect the recreational value of
natural waters by rendering them unsafe, aesthetically objectionable or otherwise unusable,
which interferes with the health, physical comfort or enjoyment on the part of the user. These
organisms are free-living species that are considered to occur naturally in recreational waters.
Guidance is provided, as both the authorities responsible for managing recreational waters and
the general public should remain informed of both the possible risks posed by these organisms
and the steps that can be taken to reduce potential exposures. This list is not intended to be
exhaustive; thus, the responsible authorities may wish to provide information on other organisms
of regional or local significance.
7.1
Schistosomes causing swimmer’s itch
Description
Swimmer’s itch (cercarial dermatitis) is caused by human reaction to dermal penetration by
parasitic flatworms or “schistosomes” that infect certain waterfowl and aquatic rodent species
(Manitoba Water Stewardship, 2007). These schistosomes belong to the family
Schistosomatidae, and species implicated as causes of swimmer’s itch include members of the
genera Austrobilharzia and Trichobilharzia (Levesque et al., 2002; CDC, 2004a). Cercarial
dermatitis should be clearly distinguished from human schistosomiasis, a far more serious human
infection that is caused by species of the genus Schistosoma and is typically restricted to tropical
regions of the world (WHO, 2003a).
The schistosomes that cause swimmer’s itch have a two-host life cycle, consisting of a primary
host (waterfowl or aquatic rodents) and an intermediate host (certain species of aquatic snails).
The adult parasite lives in the bloodstream of infected geese, swans, ducks, gulls, muskrats and
beavers and produces eggs, which are passed in the faeces of these animals. These eggs hatch in
the water, releasing small free-swimming larvae called miracidia. These larvae then seek out a
suitable snail host to infect in order to continue their life cycle. Within the snail, the parasite
further develops into a different type of larva called cercariae, which are released into the water
when conditions are appropriate. The free-swimming cercariae once again seek out a suitable
bird or animal host to begin their life cycle all over again. Humans are an accidental or dead-end
host for these organisms. If the cercariae accidentally encounter humans in the water, they may
penetrate the outer layer of the skin, but quickly die, as they cannot develop any further. The
presence of the cercariae beneath the skin causes an allergic reaction to develop (i.e., cercarial
dermatitis), which accounts for the symptoms observed in infected swimmers.
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Health effects
The effects of swimmer’s itch may be felt shortly after swimming, in some cases in as little as a
few minutes. Swimmers usually first experience a tingling, itching or burning sensation. Small,
reddish pimples typically may appear within 12 hours after infection, and these can progress to
larger secondary blisters or rashes, which can be accompanied by an even stronger itching
sensation. The infection is self-limiting, typically lasting from 2 to 5 days; however, symptoms
can persist for as long as 2 weeks. Swimmer’s itch is not contagious and cannot spread from
person to person. However, because swimmer’s itch is caused by an allergic reaction, individuals
can develop an increased sensitivity to subsequent infections. The symptoms become more
intense and develop much more rapidly in these instances (British Columbia Ministry of Health,
2005). Sensitivity can vary considerably between different individuals; some may strongly show
the effects of infection, whereas others may not show any signs of illness.
Although affected individuals are advised to seek medical treatment from a health professional,
treatments that may be effective in combatting itching include the application of cold
compresses; the use of anti-itch medications such as corticosteroid creams or calamine lotion;
and the taking of oral antihistamines or bathing in baths containing Epsom salts, baking soda or
colloidal oatmeal (British Columbia Ministry of Health, 2005; Manitoba Water Stewardship,
2007). It is recommended that affected individuals refrain from scratching, as it increases the
potential for secondary bacterial infection (CDC, 2004b).
Occurrence in the environment
These schistosomes can be encountered in fresh waters and at coastal beaches throughout
Canada and the northern United States. Indeed, swimmer’s itch has been reported in virtually
every Canadian province. Accidental introductions of host snail species are reported to have
spread this parasite from the Atlantic coast of North America to British Columbia coastal waters
(Leighton et al., 2004). Bird species (ducks and gulls) that play host to the parasite may also be
expanding in geographic range (Verbrugge et al., 2004). Reports of incidents appear to be
increasing in the United States and Canada, possibly reflecting increasing use of recreational
water bodies.
The presence of the organisms in natural waters is dependent upon a number of factors, both
biological and environmental. As a result, it is very difficult to predict when and where
swimmer’s itch might become a problem. Propagation of the organisms requires that both the
primary and intermediate host be present in sufficient numbers. As well, the species involved do
not all follow the same timetable in terms of infection of, and release from, their primary and
secondary hosts. Thus, the organisms can be encountered at different times during a recreational
water season (Michigan Department of Environmental Quality, 2005).
The cercariae are encountered in areas where the snail beds are the densest. These are typically
shallow waters, particularly those with large numbers of aquatic plants. Water temperature is
also thought to have a significant effect on the release of mature cercariae by infected snails.
Cercarial production and concentrations are thought to increase in warmer waters (Verbrugge et
al., 2004), which may partially explain why infections are encountered more frequently during
the summer months. The organisms can be carried significant distances by wind and wave
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action. Persistent onshore winds may drive them to accumulate at the shoreline, whereas
sheltered bays may act to retain the organisms within a localized area.
Related epidemiology
Much of the information on swimmer’s itch infection has come from case reports of human
illness. For most recreational waters in Canada, the risk of contracting swimmer’s itch through
recreational activity is considered to be quite low. However, many cases go unreported, as the
symptoms are typically benign and thus users may not seek out medical attention.
Levesque et al. (2002) conducted an investigation of an outbreak of cercarial dermatitis that
occurred on Lac Beauport, a recreational lake in the Quebec City region in the summer of 1999.
A case reporting form was sent to 450 families likely to have activities that would bring them
into contact with the lake’s water. Snails were characterized, and the prevalence of their
infestation by schistosomes was investigated. In total, 63 episodes consistent with cercarial
dermatitis were reported, with the symptoms affecting mainly children under 10 years of age.
Sixty-nine percent of the cases occurred from swimming at the same beach. This location was
the one where the only population of snails in the lake was identified. The people most affected
were those who bathed in shallow water along the shoreline. Mallard ducks were observed to be
present in high numbers during the 1999 summer season. Concentrations of faecal coliforms,
faecal streptococci and other bacteriological water quality indicators at the beaches were low.
Based on all of the available evidence, the authors confirmed that the cases were indeed due to
schistosomes. Shoreline residents were informed that they should not feed waterfowl, and snail
populations were reduced by removing organic wastes found within the main snail habitat. The
control mechanisms were thought to have been effective, as there were no reported instances of
cercarial dermatitis at this location during the following season.
Another study by Leighton et al. (2004) looked at case reports and the biological factors
contributing to two outbreaks of dermatitis at Crescent Beach near Surrey, B.C. Thirty-six cases
of dermatitis were reported in the summer of 2001, and 44 more cases were reported in the
summer of 2002. The clinical presentation was consistent with schistosome dermatitis or
swimmer’s itch. The causative agent was identified as the schistosome cercaria, Austrobilharzia
variglandis, carried by the introduced host snail, Ilyanassa obsoleta. Many of the cases of
swimmer’s itch appear to have occurred after exposure to the parasite in the shallow tide pools of
the upper beach where there are large aggregations of the snails and frequent human use of the
beach for wading. A survey of schistosome infections in snails demonstrated that infected snails
were present in most areas of the beach. The reasons for the sudden outbreak were unclear. Both
the snail host and the schistosome species had been known to be present at this location for
several years. Factors attributed to the sudden outbreak included increased beach use by
recreational users, seasonal environmental factors (temperature, weather), the age of the snail
population and the size of the host population.
Verbrugge et al. (2004) conducted a prospective epidemiological study to assess the incidence
and severity of swimmer’s itch among recreational water users at Douglas Lake, Michigan, in
July 2000. In total, 301 subjects were included in the analysis. Data were collected on
1300 water exposure days for the 301 swimmers, and 89 episodes of swimmer’s itch were
recorded (corresponding to an incidence of 6.8% per water exposure day). A total of 52 people
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(17.3%) experienced swimmer’s itch, with 58% of these having only one episode, 25% having
two episodes and 17% having three or more episodes. There was a highly significant association
of swimmer’s itch with shallow water use. Risk was also shown to increase with the number of
days of exposure reported. Higher incidences were observed in the southern and eastern zones of
the lake, and it was reasoned this was at least partially due to persistent onshore winds and the
existence of sheltered bays in these areas.
Managing health risks
The schistosomes responsible for causing swimmer’s itch are considered to occur naturally in
Canadian surface waters. They are not related to faecal pollution; as a result, their presence is
not indicated during standard water quality testing for the recommended indicators of faecal
contamination. The factors that are necessary for swimmer’s itch to become a problem in
recreational areas are subject to constant change. Certain areas may report a problem where none
appeared to exist previously. Similarly, areas in which swimmer’s itch has been reported will not
necessarily always remain a problem.
A management strategy combining both actions to control the extent of the water quality hazard
and steps to limit exposure during periods or in areas perceived to be of increased risk is
recommended to reduce the risk of human exposure to these schistosomes in recreational waters.
Warning signs that clearly notify the public of the risk of exposure should be posted at
recreational water areas where cases of swimmer’s itch have been reported. Additionally, a
swimming advisory may be issued at the discretion of the responsible authority. Further details
on the posting of information at recreational water areas can be found in Part I (Management of
Recreational Waters).
Another approach to risk reduction can include the provision of educational materials outlining
steps the public may take to reduce their personal risk of exposure as well as to potentially
reduce the severity of the symptoms of infection. Guidance provided in materials for public
communication may include the following recommendations:
•
Recreational water users should avoid areas where swimmer’s itch has been reported or
where signs have been posted warning users of a potential risk of infection.
•
At locations where swimmer’s itch may be suspected, users should particularly avoid areas
where schistosomes are more likely to be encountered, such as shallow waters with large
quantities of aquatic plants.
•
Users should towel down briskly upon leaving the water. Showering as soon as is practical
after recreational water activity is also recommended to help minimize the risk of infection.
•
Any user experiencing adverse health effects from recreational water activity should consult
a medical professional and, if necessary, alert the appropriate authorities.
•
Users are also reminded not to feed any waterfowl that may be present, as these animals can
harbour the organisms responsible for causing swimmer’s itch.
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7.2
Aquatic vascular plants and algae
Description
Aquatic vascular plants (macrophytes) and algae can affect recreational water uses. It is difficult
to estimate the magnitude of the adverse effects of these organisms in terms of their degree of
interference with recreational pursuits or the potential risks to health posed to recreational water
users.
The presence of these organisms can present a safety risk to recreational water users. Swimmers
may become entangled in the fronds of aquatic plants. Dense growths can obstruct the view of
the bottom and underwater hazards and may impair the ability of safety personnel to see persons
in distress. Algal matrices attached to rocks and other substrata (i.e., periphyton) can cause
slippery conditions that may lead to unintentional immersions or injuries.
Excessive growth of the organisms can also create aesthetic problems for recreational water
areas. Macrophytes can assume high population densities and make nearshore and shallow
regions unsuitable for any purpose (Priyadarshi, 2005). Dislodged rafts or mats of plants and
algal material can be washed ashore and left to rot, fouling beaches. In addition to being
unsightly, these masses can further detract from user enjoyment by producing offensive odours
and restricting access to shorelines. They may also pose a public health risk, as they can attract
undesirable animals to the area and provide breeding grounds for a variety of species of insects
and bacteria (Whitman et al., 2003). The most notorious organism in this respect has been the
green algal species Cladophora (Priyadarshi, 2005). There have been countless accounts of
beaches and shorelines fouled by rotting, stinking masses of this alga. Recent research has
indicated that Cladophora mats can provide a secondary habitat for bacteria that could
potentially influence water quality in affected swimming areas (Whitman et al., 2003; Ishii et al.,
2006b).
Increased plant growth (both macrophytes and algae) can be caused by the presence of excess
nutrients. Various nutrient sources, including agricultural practices, domestic sewage and
industrial effluent, all increase the amount of phosphorus and nitrogen in aquatic systems. The
result of this increase in nutrients is called cultural eutrophication. Impaired water quality as a
result of high algal populations and eutrophicating conditions can reduce recreational
opportunities (Chambers et al., 2001). Canadian Water Quality Guidelines have been developed
for both phosphorus and nitrogen species to protect the aquatic environment from nutrients and
their effects on aquatic organisms (CCME, 1999).
Managing health risks
Excessive numbers of aquatic plants and algae should be absent from areas intended for use as
swimming beaches. Similarly, recreational activities should not be pursued in areas where these
organisms are present in quantities such that the responsible authorities deem that their presence
poses a potential health or safety risk to recreational water users. It is recommended that an
Environmental Health and Safety Survey be conducted at the start of each swimming season to
identify potential safety hazards that may be encountered at a given recreational water area.
Subsequently, one barrier to risk may involve the posting of signs warning users of potential
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visibility or entanglement risks that may be posed by these organisms. Further information on the
posting of warning signs can be found in Part I (Management of Recreational Waters).
Many aquatic plants and algae also provide an important habitat for fish and other aquatic biota.
Management actions that try to remove these oplant rganisms from natural waters are
discouraged, as removal may be harmful to the aquatic environment and is generally not
effective from both a practical (plants quickly repopulate) and economical perspective. The
application of pesticides to combat these organisms is similarly not recommended, as their use
may create a health hazard for recreational water users if used incorrectly, and they are also
detrimental to the healthy functioning of the aquatic ecosystem.
Improved beach cleanup procedures to remove masses of plants and algal material that may have
washed up on shorelines represents another barrier that can be effective in reducing potential
risks to recreational users. Longer-term actions that may also be considered as barriers to reduce
the impact of these organisms can involve the identification of the major nutrient inputs within
the watershed and the subsequent implementation of strategies for their control.
7.3
Additional organisms
Numerous other organisms can interfere with the safe and enjoyable use of recreational waters in
Canada. For example, at some coastal beaches, jellyfish can cause painful and possibly serious
stings for recreational water users who come in contact with them. Similarly, leech “bites” can
occur in leech-infested areas, and sea urchins and mussel shells can cause painful injuries when
stepped on by users of recreational waters. As these organisms are often of local or regional
significance, it is recommended that, where necessary, the responsible authority provide the
appropriate guidance to recreational water users on these subjects. This may include providing
information on the potential risks posed by these organisms as well as steps for reducing the
public’s individual risk of exposure.
Summary
1. Recreational water activities should not be pursued in areas where any of the organisms
described in this section are known to be present in quantities such that they are deemed to
pose a potential health or safety risk to recreational water users. An Environmental Health
and Safety Survey is an important tool for identifying water quality hazards that may be
encountered within the recreational water area.
2. Effective barriers to protect users from safety risks that may be encountered during
recreational activities can include the provision of warning signs or educational materials
informing users of the potential risks posed by a given water quality hazard, steps to reduce
their risk of contact and possible actions to be taken in the event that exposure does occur.
8.0
Physical, aesthetic and chemical characteristics
This section describes the major physical, aesthetic and chemical characteristics of water that
may affect recreational water bodies. Information is provided as to the potential effects each may
have on the safe, enjoyable use of recreational waters. Guideline values or aesthetic objectives
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have been provided where possible. It is intended that the values and associated guidance
apply to all recreational waters, regardless of the types of activities practised. Responsible
authorities may wish to establish separate guideline values or aesthetic objectives for waters not
intended for primary contact use at their discretion.
Methods for determining physical, aesthetic and chemical characteristics of recreational waters
can be found in Standard Methods for the Examination of Water and Wastewater (APHA et al.,
2005). Sampling for these parameters is at the discretion of the responsible authority, however it
is suggested that valuable times may include:
• when conducting the EHSS;
• at the start of, and at regular intervals during the swimming season,
• during area assessments conducted in response to recreational water issues, as appropriate.
8.1
Physical characteristics
8.1.1
pH
Guideline values
Both alkaline and acidic waters may cause eye irritation. To be protective against the risk of eye
irritation, the pH of recreational waters should be in the range of 5.0 to 9.0.
Description
Mood (1968) concluded that exposure to water is foreign to the eye and may, under certain
circumstances, be very irritating. He assumed that the ideal, non-irritating solution would have
the same physicochemical properties as tears, including a pH of 7.4, although there is some
evidence to suggest that ophthalmological solutions slightly more alkaline are actually preferred
(Raber and Breslin, 1978).
Mood (1968) reported that tears have the capacity to rapidly neutralize an unbuffered solution
from a pH as low as 3.5 or as high as 10.5. The neutralizing capacity of the tears would be
exceeded by highly buffered waters. However, Mood (1968) concluded that unbuffered waters
are not found in nature under normal conditions; hence, he suggested that the pH range for water
with low buffering capacity should be between 5.0 and 9.0. Dillon et al. (1978) reported that
most lakes in south-central Ontario have 10–200 µeq/L of acid-neutralizing capacity (ANC), and
many of these lakes have depressed pH.
Studies completed by Basu et al. (1984) used water from two inland lakes in Ontario: Clearwater
Lake (pH approximately 4.5), with an ANC of −40 µeq/L (Ontario Ministry of the Environment,
1980), and Red Chalk Lake (pH approximately 6.5), with an ANC of 70 µeq/L. The eyes of both
test rabbits and human volunteers were exposed to these waters, and no significant differences
were observed in their reactions (Basu et al., 1984). Basu et al. (1984) exposed one eye to the
low-pH water and the other to the higher-pH water. The human eyes were exposed for 5-minute
periods, and no ocular effects were noted. The rabbit eyes were exposed for periods of 15
minutes and checked for ocular reactions in terms of conjunctival congestion, corneal epithelial
staining with fluorescein, epithelial cell and leukocyte content of tears, change in tear molarity
and the penetration of fluorescein into the anterior chamber. Basu et al. (1984) concluded that the
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exposure of healthy eyes to lake water having a pH as low as 4.5 is not harmful to the external
ocular tissues.
8.1.2
Temperature
Guideline value
Precise guideline values for the temperature of waters to be used for swimming cannot be
established. Tolerance to water temperatures can vary considerably from individual to individual.
Users should not engage in recreational activities at temperature-time combinations sufficient to
cause an appreciable increase or decrease in their core body temperature.
Description
Cold water exposures
Water is very efficient at conducting heat away from the body. When in water, unlike air, the
surface area available for heat exchange reaches close to 100% (Transport Canada, 2003). Water
has 25 times the thermal conductivity of air and cools the body 4–5 times faster than air of the
same temperature (Tipton and Golden, 2006).
The definition of cold water must be considered with respect to normal body temperature,
duration of exposure and degree of protection by insulation (Canadian Red Cross, 2006). When
heat loss exceeds heat production, the temperature of the body will drop below the normal value
of 37°C (Tipton and Golden, 2006). Water, even at comfortable levels, can result in the transfer
of heat away from the body. Fatalities from cold water immersion have been reported in
subtropical waters in which exposure was prolonged. Thermal neutrality in water is reported to
occur at 35°C. Below this value, the human body is expected to lose more heat than it is capable
of producing. Sudden, unprotected immersion in water of ≤ 15°C is considered to be a
potentially life-threatening situation (Canadian Red Cross, 2006).
Figure 2. Relationship between water temperature and survival time in cold water (adapted from the Royal Life
Saving Society of Canada, 1976).
Experts have identified four sequential stages that can occur following immersion in cold water:
1) gasping and cold shock, 2) swimming failure, 3) hypothermia and 4) post-rescue collapse
(Transport Canada, 2003; Canadian Red Cross, 2006). It is believed that most deaths in cold
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water occur from drowning by submersion of the airways during the first two stages of cold
water immersion (Canadian Red Cross, 2006).
The rate of body cooling and the incidence of survival in cold water can vary considerably from
individual to individual. This variability can be related to age, sex, body size, ratio of body mass
to surface area, percentage body fat and overall physical fitness. The ratio of body mass to
surface area is greater in large, heavy individuals, and their temperatures change more slowly
than those of small children (Kreider, 1964). Other factors influencing cooling can include the
degree of protective clothing and physical behaviour and body posture in the water. The use of
drugs or alcohol can also exacerbate the effects of cold water immersion, by impairing alertness
and motor skill use and by interfering with the body’s temperature regulation mechanisms
(Canadian Red Cross, 2006).
Immersion in cold water may occur through intentional or unintentional activities. Persons using
recreational waters should remain aware of the risks involved and take appropriate precautions
against cold water exposure. The Canadian Red Cross (2008) and Transport Canada (2006) have
produced publications that provide information on survival in cold waters. Proper protective
garments such as a wetsuit or survival suit should be worn during recreational water activities
where cold water exposure is anticipated. Similarly, precautions should be taken against
accidental immersions, including use of safety lines and wearing of proper personal flotation
devices.
Warm water exposures
By comparison, relatively little information is available regarding the physiological effects
of human exposure to warm waters. Early communication on this topic suggested that,
physiologically, neither adults nor children would experience thermal stress under modest
metabolic heat production, as long as the water temperature was lower than the normal skin
temperature of 33°C (Newburgh, 1949). Water ranging in temperature from 20°C to 30°C is
considered comfortable for most swimmers (WHO, 2003a, 2006).
In Canada, under most circumstances, ambient temperatures observed during the summer months
do not reach levels sufficient to elevate recreational water temperatures above normal human
body temperature. Natural hot springs—thermal springs that can reach temperatures in excess of
37°C—are a notable exception. Individuals using these types of facilities should monitor their
exposures carefully so as to avoid overheating. For the majority of recreational water areas, the
heat effects observed during summertime water activities are largely attributed to sun exposure.
Numerous health authorities have provided guidance on avoiding heat exposure during outdoor
activities; thus, this information can be viewed as extending to recreational water exposures as
well.
Heat stress or heat exhaustion can occur following vigorous exercise in warm environments.
Signs of heat exhaustion can include excessive sweating, elevated temperature or pulse rate,
headache and dizziness or weakness. The Canadian Red Cross has similarly developed
publications on the prevention of heat-related illness or injury (Canadian Red Cross, 2011).
Precautionary measures to minimize the effects of heat exposure during recreational water
activities are similar to those for reducing sun exposure. These include wearing lightweight
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clothing and broad-brimmed hats, seeking out cool or shady areas, avoiding activity during
midday periods when the sun is most intense, ensuring an adequate supply of drinking water and
replenishing any salt losses.
8.2
Aesthetic characteristics
Good aesthetic quality is an important consideration in ensuring the maximum use and
enjoyment of recreational waters. A recreational water area should be considered aesthetically
acceptable to its users. Waters used for recreation should be free from substances (either
attributable to human activities or due to natural processes) that impair its aesthetic appreciation.
These can include:
•
•
•
•
substances producing objectionable colour, odour, taste or turbidity;
floating debris, oil, scum and other matter;
materials that will settle to form objectionable deposits; and
substances and conditions or combinations thereof in concentrations that produce undesirable
aquatic life.
The term aesthetic has been defined as “concerned with beauty or the appreciation of beauty”
(Canadian Oxford Dictionary, 2004). Thus, not only should a recreational water area be free
from objectionable factors, but various other aesthetic components of the aquatic ecosystem and
surrounding land should be present. Recreational waters should also be considered free from
substances in amounts that would interfere with the existence of life forms of aesthetic value.
This section discusses parameters that may affect the aesthetic quality of a recreational water
area. For the purposes of this document, it is the effects that these factors may have on aesthetic
perception that are of primary significance. However, it should be noted that with these
parameters there also exists certain implications for human health and safety. For example,
waters in which the visibility has become significantly impaired can present a safety risk for
recreational water users.
Aesthetic objectives for turbidity, clarity and colour have been proposed; however, it is
recognized that the natural levels of these parameters in Canadian waters can vary considerably.
Thus, it is recommended that, when evaluating these parameters as part of an Environmental
Health and Safety Survey, the associated values also be interpreted as not being significantly
increased over that which would be considered natural background.
8.2.1
Turbidity
Aesthetic objective
An aesthetic objective of 50 nephelometric turbidity units (NTU) is suggested for recreational
waters.
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Description
Standard Methods for the Examination of Water and Wastewater (APHA et al., 2005) defines
turbidity as an “expression of the optical property that causes light to be scattered and absorbed
rather than transmitted with no change in direction or flux level through the sample.” The current
method of choice for turbidity measurements in Canada is the nephelometric method, and the
unit of turbidity measured using this method is the nephelometric turbidity unit, or NTU (Health
Canada, 2012c).
Turbidity in water is caused by suspended and colloidal matter, such as clay, silt, finely divided
organic and inorganic matter, plankton and other microscopic organisms (APHA et al., 2005).
This parameter is important for aesthetic, safety and, to a lesser degree, health reasons. High
turbidity is aesthetically displeasing and can be a safety concern when it reduces visibility
through the water. Because filtration equipment and modern water treatment processes are not
feasible at natural swimming areas, safety concerns associated with turbid or unclear water are
dependent upon the intrinsic quality of the water itself. Lifeguards and other persons near the
water must be able to see and distinguish people in distress. In addition, swimmers should be
able to see quite well while under water.
Health considerations associated with turbidity are related primarily to the ability of particles to
adsorb microorganisms and chemical contaminants. This can have a number of important effects
on water quality:
•
Suspended particles can provide a measure of protection for microorganisms (bacteria,
viruses, protozoa) that have been adsorbed to their surface by shielding them from the effects
of environmental stresses such as UV radiation and predation by higher microorganisms.
•
The presence of turbidity may interfere with the quantification of faecal indicator organisms.
In the enumeration of bacteria, it is assumed that each colony represents one cell; however, a
single colony could emanate from a particle containing many bacterial cells adsorbed on its
surface. Fewer cells than were actually present would then be recorded. This phenomenon
would also lead to an underestimation of bacterial numbers with the MPN technique (Health
Canada, 2012c).
•
Particulate matter may also contain chemical contaminants such as heavy metals and biocides
(Health Canada, 2012c).
Surface water levels can vary from 1 to more than 1000 NTU (Health Canada, 2012c). Runoff
water quality measurements indicated levels of 4.8–130 NTU during the first hour of an urban
rainfall occurrence (U.S. EPA, 1978). In the quiescent zone of a swimming beach or other
recreational water area, it is suggested that turbidity measurements in the vicinity of 50 NTU
would be sufficient to satisfy most recreational uses, including swimming.
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8.2.2
Clarity (light penetration)
Aesthetic bbjective
Water should be sufficiently clear that a Secchi disc is visible at a minimum depth of 1.2 m.
Description
The Canadian Oxford Dictionary (2004) defines “clarity” as “the state or quality of being clear.”
Clarity is associated with the distance light can penetrate into a body of water and is often
interpreted as “how far can one see into the water?” The clarity of a water body can be simply
evaluated by using a Secchi disc, which is a device used to approximately measure visibility
depths in water. The upper surface of a circular metal plate, 20 cm in diameter, is divided into
four quadrants and painted so that the two quadrants directly opposite each other are black and
the intervening ones are white. When suspended to various depths of water by means of a
graduated line, its point of disappearance indicates the limit of visibility. It is then raised until it
reappears, and the average of the two depths is taken as the Secchi disc transparency.
The principal factors affecting the depth of light penetration in natural waters include suspended
microscopic and macroscopic organisms, suspended mineral particles, stains that impart a colour,
detergent foams, dense mats of floating and suspended debris, or a combination of these factors.
It is important that water at swimming areas be clear enough for users to estimate depth, to see
subsurface objects easily and to detect the submerged bodies of swimmers or divers who may be
in distress. Aside from the safety factor, clear water fosters enjoyment of the aquatic
environment.
For primary contact recreation waters, it has been suggested that clarity be such that a Secchi
disc is visible at a minimum depth of 1.2 m (Environment Canada, 1972). In “learn to swim”
areas, the clarity should be such that a Secchi disc on the bottom is visible. In diving areas, the
clarity shall equal the minimum required safety standards, depending on the height of the diving
platform or board (National Technical Advisory Committee, 1968).
8.2.3
Colour
Aesthetic objective
No numerical value can be established for colour in recreational waters. Colour should not be so
intense as to impede visibility in areas used for swimming.
Description
The observed colour of water is the result of light backscattered upward from a water body after
it has passed through the water to various depths and undergone selective absorption (CCME,
1999). There are two measures of colour in water: true and apparent. The term colour is most
often used to mean true colour, the colour of water from which turbidity has been removed. To
measure true colour, the water has to be filtered or centrifuged to remove the sources of apparent
colour. The standard method for measuring colour is the platinum-cobalt method (APHA et al.,
2005). Colour is measured by visual comparison with coloured solutions of known
concentrations. Under this method, one Pt-Co unit is equivalent to that produced by 1 mg
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platinum/L in the form of the chloroplatinate ion. The ratio of cobalt to platinum given matches
the true colour of water.
The correct units for true colour are colour units (CU), with one colour unit being equivalent to
1 Pt-Co unit (APHA et al., 2005). True colour can range from less than 5 CU (colour units) in
very clear waters to 1200 CU in dark peaty waters (Kullberg, 1992). Natural minerals give true
colour to water; for example, calcium carbonate in limestone regions gives a greenish colour,
whereas ferric hydroxide gives a red colour. Organic substances, tannin, lignin and humic acids
from decaying vegetation also give true colour to water (Department of National Health and
Welfare, 1979).
Apparent colour includes not only colour due to substances in solution but also that due to
suspended matter (APHA et al., 2005). Measurements for apparent colour are determined on
the original sample without filtration or centrifugation (APHA et al., 2005). Apparent colour is
usually the result of the presence of coloured particulates, the interplay of light on suspended
particles and such factors as bottom or sky reflection. An abundance of cyanobacteria can impart
a dark greenish hue to water, whereas diatoms or dinoflagellates may give a yellowish or yellowbrown colour. There are algae that impart a red colour, and, rarely, zooplankton, particularly
microcrustaceans, may tint the water red. Polluted waters may have strong apparent colour.
Industrial discharges (particularly those from the pulp and paper and textile industries) can be
highly coloured and thus may affect water coloration. Factors increasing the turbidity of natural
waters may similarly affect apparent colour.
Colour in lakes may not be uniform from surface to bottom; also, the colour may change
periodically. Increases in surface runoff contribute great quantities of inorganic and organic
substances. Summer or early autumn production of phytoplankton blooms causes lakes to
become a “soupy green,” which disappears later in the season. Exposure to light causes
bleaching of certain colours in natural waters, and this effect will vary according to transparency.
Colour may also be dependent on factors that affect the solubility and stability of the dissolved
and particulate fractions of water, such as temperature and pH.
Generally, a rich, highly productive lake may appear yellow, grey-blue or brown as a result of
quantities of organic matter, and less productive lakes tend towards blue or green caused by
differential light absorption and scattering of different wavelengths (Ruttner, 1963; Reid and
Wood, 1976).
The causes of colour in marine waters are not thoroughly understood, but dissolved substances
are one of the contributory factors. The blue of the sea is a result of the scattering of light by
water molecules, as in inland waters. Suspended detritus and living organisms give colours
ranging from brown through red and green. Estuarine waters are not as brilliantly coloured as the
open sea; the darker colours result from the high turbidity usually found in such situations (Reid
and Wood, 1976).
The main effects of water colour on recreational activities are aesthetic and safety related. Very
dark water restricts visibility both for swimmers and for people concerned with their safety. In
recreational waters, it is desirable that the natural colour of the water not be altered by any
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human activities. A maximum of 100 CU was proposed by Environment Canada (1972), while a
guideline of not more than 30 CU above natural value is used by Alberta Environment (1999).
Supporting evidence for these values has not been given. Thus, in the absence of strong
supporting evidence, it is recommended that they be considered as guidance values.
8.2.4
Oil and grease
Aesthetic objective
No numerical value can be established for oil and grease in recreational waters. Oil, grease or
petrochemicals should not be present in concentrations that:
•
•
•
can be detected as a visible film, sheen or discoloration on the surface;
can be detected by odour; or
can form deposits on shorelines and bottom sediments that are detectable by sight or odour
(International Joint Commission, 1987).
Description
Standard Methods for the Examination of Water and Wastewater (APHA et al., 2005) defines
“oil and grease” as “any material recovered as a substance soluble in the solvent.” The category
of oil and grease includes many different substances of mineral, animal, vegetable or synthetic
origin—all of which can have vastly different physical, chemical and toxicological properties.
Consequently, it is very difficult to establish a numerical criterion for oil and grease.
Contamination of recreational waters with oily substances may have natural origins or may be
a result of human activities. Some oils are of natural origin, such as seepage from natural
underwater oil deposits or from the decomposition of some materials. For example, natural
biological populations release lipid compounds, which can form natural slicks.
Human-made contamination is of greatest concern. It can come from a number of sources, such
as the discharge of industrial wastes, road runoff, residual hydrocarbon deposits from motorboat
engine exhaust emissions, the discharge of fuel tank contents of ships, either accidentally or
deliberately, and shipwrecks. Marinas and boat launches can also be important sources of oil and
grease contamination for recreational waters.
Very small quantities of oily substances make water aesthetically unattractive. Oils can form
films, and some volatile components may create odours or impart a taste to water (WHO, 2003a).
Oil and grease may foul equipment, shorelines or the bodies of swimmers. The possibility exists
that recreational users might still use the water in cases of low contamination. The risk of
toxicity from exposure to oily substances through ingestion, skin absorption or inhalation of
vapours during recreational water activity is regarded to be low. Oils and greases of animal or
vegetable origin are generally considered to be non-toxic to humans. Similarly, it is recognized
that petroleum compounds become organoleptically objectionable at concentrations far below the
levels required for chronic human toxicity. Thus, the consumption of oil-polluted water is
unlikely to be a significant source of exposure for humans (Train, 1979).
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8.2.5
Litter
Aesthetic objective
No numerical value can be established for litter at recreational water areas. Recreational water
areas should be free from floating debris as well as materials that will settle to form
objectionable deposits.
Description
There are a variety of types of litter that can be found in recreational waters or deposited on
beaches. Some examples include discarded food and food packaging, cardboard and paper
products, plastic containers, styrofoam materials, rubber goods, aluminium cans, broken glass,
discarded clothing, cigarette butts, medical wastes and dead animals. Large accumulations of
seaweed and algae are not technically litter, but they are likely to pose an aesthetic problem (both
visually and due to potential odour).
In addition to being aesthetically undesirable, the presence of litter can also present a health and
safety risk to recreational water users. Some materials can be injurious to recreational water
users who come in direct contact with them. Discarded litter also has the potential to attract
wildlife, which can contribute to the faecal contamination of recreational waters. Indeed, litter
counts have been considered as a possible indicator of the potential of acquiring gastrointestinal
illness though recreational water activities. Similarly, flying and/or biting insects may also be
associated with litter. These are considered at the very least a nuisance and could potentially pose
a health threat in the form of zoonotic disease (NHMRC, 2008).
8.3
Chemical characteristics
Guideline values
There is insufficient information to support the establishment of guideline values for specific
chemical parameters in recreational waters. Risks associated with specific chemical water
quality hazards will be dependent on the particular circumstances of the area in question and
should be assessed on a case-by-case basis.
In general, potential risks from exposure to chemical parameters will be much smaller than the
risks from the microbiological hazards potentially present in recreational waters (WHO, 2003a).
With chemical concentrations typically found in water, most recreational water users will not be
exposed to sufficient concentrations necessary to elicit either an acute or chronic illness
response.
Description
Chemical contaminants can enter recreational waters or be deposited on beaches from both
natural and anthropogenic sources (WHO, 2003a). These include point sources, such as
industrial outfalls or natural springs, and non-point (diffuse) sources, such as runoff from urban
or agricultural areas.
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Inorganic chemicals
National surveys of the water quality of lakes and rivers used for recreational activities indicate
that concentrations of inorganic chemicals are low (NAQUADAT, 1988; Government of
Canada, 1991). Analyses for heavy metals indicated that they are present in concentrations
considerably below those recommended as guidelines for drinking water quality (Government of
Canada, 1991). Ingestion would be considered the primary pathway of exposure for inorganic
chemical contaminants; however, skin absorption is recognized as a route of uptake for certain
heavy metal forms. Owing to the low concentrations encountered in most
natural waters and the types and degrees of exposure involved during typical recreational water
activities, exposure to inorganic chemical contaminants is not considered to represent a
significant health risk for recreational water users at recognized swimming areas.
Organic chemicals
There are many sources of contamination by organic chemicals, including industrial manufacture
and use and domestic use of such items as paints, fuels, dyes, glues, pesticides and cleaning
supplies (NAQUADAT, 1988; Health Canada, 1997).
National surveys have analysed the level of contamination of recreational waters by organic
chemicals. The concentrations of organic chemicals that have been detected in waters that could
be used for recreational purposes were lower than the recommended drinking water guidelines
(Government of Canada, 1991; Marvin et al., 2004) and thus should not pose a significant threat
to human health.
It has been suggested that for some organic chemicals, skin absorption can be as important as
ingestion in contributing to exposure (Brown et al., 1984; Moody and Chu, 1995). However, it is
generally concluded that given the low concentration of organic contaminants encountered in
most natural waters and the typical exposure scenarios encountered during recreational water
activities, it is not likely that dermal exposure presents a significant risk (Moody and Chu, 1995;
Hussain et al., 1998). Nonetheless, precautionary measures such as restricting swimming to
public beaches and showering with soap and water following recreational activity will further
ensure that any risk is minimized.
Managing health risks
The risk of human exposure to chemical contaminants in Canadian waters through recreational
activities is considered low. Nevertheless, scenarios do exist that may contribute to the presence
of a chemical water quality hazard at a particular recreational water body. As a result, it is
important for beach operators and service providers to have a mechanism in place to ensure that
potential chemical hazards and their risks are recognized. An Environmental Health and Safety
Survey is an important tool for helping recreational water area operators identify and assess
potential sources of chemical contamination that are relevant to their beach area.
Risks associated with specific chemical water quality hazards will be dependent on the particular
circumstances of the area under consideration. Thus, in all instances, the risk of human exposure
to chemical contaminants in recreational waters must be assessed on a case-by-case basis, taking
local factors into account. In general, some key elements that should be included in any approach
to assessing chemical water quality hazards in recreational waters are as follows:
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•
•
•
•
•
historical understanding of the area to identify past activities that may result in contaminated
water and/or sediments;
inspection of the recreational water area to identify any obvious sources of chemical
contamination, such as outfalls or discharges;
conducting of additional steps as necessary to support a quantitative health risk assessment,
including chemical analysis of representative samples (using methods deemed acceptable by
the regulating agencies) and review of the available toxicological information on the
chemical contaminant(s) in question;
consideration of the pattern and type of recreational activity to determine whether significant
pathways of human exposure exist (e.g., through ingestion, inhalation or skin absorption);
and
consideration of the effects of the water body dimensions (area, depth) and other
hydrodynamic and meteorological characteristics (tides, currents, prevailing winds) on the
impact of the chemical water quality hazard in question.
A multi-barrier approach represents the most effective way of protecting recreational water users
from the risk of exposure to chemical contamination at recreational water areas. This approach
uses an Environmental Health and Safety Survey to highlight potential chemical water quality
hazards as well as to identify barriers that may be implemented to both reduce the risk of
chemical contamination and restrict swimmer exposures during periods or in areas perceived to
be of increased risk.
9.0
Faecal contamination and beach sand
This section provides information on the issue of faecal contamination and faecal indicator
bacteria in beach sand, including the effects on recreational waters and steps that can be taken
to reduce swimmer exposures to contamination in the sand environment.
Currently, there is no conclusive evidence of a relationship between contact with beach sand
and illness among recreational water users, and no guideline values can be established for
concentrations of the recommended indicators for recreational waters for faecal contamination in
beach sand. Routine monitoring of sand samples for the presence of faecal indicators would not
be considered practical and is thus not recommended. Certain circumstances may warrant testing
of sand and sediment samples, such as during investigations of potential waterborne disease
outbreaks or when conducting an Environmental Health and Safety Survey.
Further research is needed to more fully characterize the relationships between faecal indicator
bacteria and the possible presence of faecal pathogens in beach sand, as well as the potential
implications to human health. Combining actions, procedures and tools to collectively reduce the
risk of swimmer exposure to faecal contamination in beach sand and recreational waters
represents the most effective approach to protecting the health of recreational water users.
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Description
Microorganisms are a recognized natural component of beach sand, with numerous species of
bacteria, viruses, parasites and fungi all having been isolated from the sand environment (WHO,
2003a). The results of recently published studies demonstrating that faecal indicator bacteria can
be isolated in high numbers from foreshore and nearshore beach sand and sediments (Alm et al.,
2003; Whitman and Nevers, 2003; Edge and Hill, 2007) have prompted renewed interest in the
question whether sand can serve as a vehicle for the transmission of pathogenic microorganisms
to beach users. Understandably, this issue is one of increasing concern to beach operators, public
health officials and the beach-going public. Beach users often spend more time on the beach than
in the water, and children routinely play in the sand at the water’s edge.
Occurrence of faecal indicator bacteria in beach sand
Enteric bacteria are not expected to survive well once displaced from their primary habitat (the
intestinal tract of humans and animals) to a secondary habitat such as the aquatic environment
(Winfield and Groisman, 2003; Anderson et al., 2005). Their survival in this environment is
influenced by a complex array of biological and environmental factors. Selective pressures that
can have a negative effect on the survival of microorganisms in recreational waters include
sunlight; osmotic stress; large variations in temperature, pH and salinity; low nutrient
availability; and competition and predation by other microorganisms (Winfield and Groisman,
2003).
Beach sand and similar environments (foreshore and nearshore soils and sediments) can provide
more favourable conditions for survival of microorganisms than the adjacent waters. Whitman
and Nevers (2003) suggested that beach sand provides a suboptimal, yet viable, environment for
the survival of enteric bacteria by providing protection from sunlight, buffered temperatures, a
degree of cover from predation, large surface areas for biofilm development and a replenishing
supply of moisture and organic nutrients from wave swash.
Numerous authors have observed the presence of faecal indicator bacteria in beach sands and
sediments, often at concentrations several orders of magnitude higher than those of the adjacent
swimming waters. At a Lake Michigan beach, Whitman and Nevers (2003) detected E. coli
counts in foreshore and submerged sand (1000–10 000 mean cfu/100 mL) that exceeded those of
the swimming water (100–1000 mean cfu/100 mL). Similar findings were reported by
Williamson et al. (2004) in a survey of E. coli densities in swimming water and sand pore water
at Lake Winnipeg beaches. Alm et al. (2003) reported that, on average, concentrations of E. coli
and enterococci were 3–17 times and 4–38 times greater, respectively, than in the water column
at beaches located on Lake Huron and the St. Clair River. Edge and Hill (2007) found E. coli
concentrations as high as 114 000 cfu/g dry sand at a Lake Ontario beach.
Similar studies have shown that faecal indicator bacteria can be isolated from other habitats
within a beach watershed, such as backshore sand (Byappanahalli et al., 2006), subtropical and
temperate stream sediments (Byappanahalli et al., 2003; Jamieson et al., 2003, 2004; Ferguson et
al., 2005; Ishii et al., 2006a), temperate forest soils (Byappanahalli et al., 2006) and within mats
of the green algal species Cladophora (Whitman et al., 2003; Ishii et al., 2006b).
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A concept that is receiving increasing attention within the scientific community is that enteric
bacteria may be capable of multiplying in the sand and similar environments in tropical and
subtropical climates provided that the proper conditions for growth are met (Davies et al., 1995;
Byappanhalli and Fujioka, 1998; Solo-Gabrielle et al., 2000; Desmarais et al., 2002; Anderson et
al., 2005). These would include (among other factors) suitable conditions of temperature,
moisture and available nutrients, as well as reduced competition from other microflora. Warmer
temperatures and higher soil nutrient concentrations have been cited as factors that may favour
the multiplication of indicator bacteria in tropical soils (Hardina and Fujioka, 1991; Whitman
and Nevers, 2003). Whitman and Nevers (2003) proposed that certain conditions that exist at
tropical beaches may be encountered at temperate beaches in the United States during the
summer months.
During an investigation of sand and water quality at a Lake Michigan beach, Whitman and
Nevers (2003) observed that in freshly replaced beach sand E. coli was able to recolonize to
earlier levels within a period of 2 weeks. The authors were not able to confirm whether the effect
was due to multiplication in an unexploited environment or the result of deposition by external
sources. Kinzelman et al. (2004a) detected similar levels of diversity among E. coli strains
collected from water, foreshore sands and submerged sands at a Racine, Wisconsin, beach. It was
suggested that accumulation, and not replication, was the major contributor to E. coli observed in
the sand.
A few researchers have suggested that some portion of the sand- or soilborne strains of E. coli
may represent a genetically distinct group, separate from the majority of isolates that dominate
in the host sources (Winfield and Groisman, 2003; Byappanahalli et al., 2006; Edge and Hill,
2007). Moreover, it is proposed that these strains may have adapted for prolonged survival and
possibly growth in the soil environment (Winfield and Groisman, 2003; Byappanahalli et al.,
2006). Byappanahalli et al. (2006) reported finding evidence of a genetically diverse E. coli
population among forest soils located within a Lake Michigan watershed. Ishii et al. (2006a)
reported finding evidence of “naturalized” E. coli populations in northern temperate soils located
within Lake Superior watersheds. The authors further reported that these strains were capable of
multiplying in non-amended, non-sterile soils at temperatures at or above 30°C.
Effects of beach sand on microbiological water quality
There has been a wealth of information in support of the notion that beach sand can present a
significant non-point source of faecal contamination for swimming waters (Alm et al., 2003;
Whitman and Nevers, 2003; Williamson et al., 2004). Faecal indicator bacteria in sand can
originate from a variety of faecal pollution sources. Gulls in particular are thought to be a
significant source of faecal contamination for beaches (Levesque et al., 1993; Fogarty et al.,
2003; Williamson et al., 2004). Canada geese populations can also present a source of faecal
contamination at areas adjacent to surface waters (Alderisio and DeLuca, 1999).
Mechanisms of transfer of sandborne contamination to the water environment include wave
swash, rain-mediated runoff and direct transfer from swimmers. Resuspension of nearshore
sediments can occur through a number of mechanisms, including wave action (including those
artificially generated by commercial and recreational boating), storms and swimmer activities.
Boehm et al. (2004) proposed that the wave-driven and tidally driven recirculation of water
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through the beach aquifier may also present a mechanism for transfer of microorganisms and
nutrients from the sand environment to swimming waters.
Pathogenic microorganisms in beach sand
Few studies have been published regarding the occurrence and survival of enteric pathogens in
beach sand and sediments. Bolton et al. (1999) reported that culturable Campylobacter and
Salmonella species were detected in 45% and 6% of sand samples, respectively, in beach sand
samples collected from various UK coastal beaches. Obiri-Danso and Jones (1999) reported
that Campylobacter could be detected year-round in low numbers (< 0.5 log cfu/g dry weight) in
river sediments at two freshwater swimming sites in the northwest of England. Salmonella was
not detected at either site. In a follow-up study at coastal beaches within the same watershed,
Obiri-Danso and Jones (2000) detected Campylobacter (C. lari, urease-positive thermophilic
campylobacters) in sediment samples collected during the winter months only. C. jejuni and C.
coli could not be isolated, and Salmonella was not detected at any point during the analysis. The
authors concluded that the sediments were not acting as a reservoir for these pathogens in this
system. NWRI (2006) found Campylobacter species commonly occurring in sand pore water at
two bird-contaminated beaches in Hamilton Harbour, Lake Ontario. C. jejuni were more
common than C. coli or C. lari, although all species appeared to be in low numbers in the pore
water.
Others have reported the detection of S. aureus and P. aeruginosa in beach sand (Papadakis et
al., 1997; Esiobu et al., 2004). P. aeruginosa is a relatively hardy species of bacteria that is
widely distributed in the aquatic environment and is known to cause skin rashes and eye and ear
infections in swimmers. Humans are the primary source of S. aureus in recreational waters; thus,
its presence in beach sand is thought to be directly related to swimmer activities. The organism is
known to be associated with skin infections in swimmers (rashes, infected cuts and scratches).
Very few data are available regarding the presence of other potential recreational waterborne
pathogens (viruses, protozoan parasites) in beach sand (WHO, 2003a).
Related epidemiology
Despite the findings that faecal indicator bacteria and potentially pathogenic microorganisms can
be detected in beach sand, there has been little evidence published to indicate a link to illness
among beach users. Marino et al. (1995) reported that there was no evidence of a relationship
between the incidence of skin symptoms and sand concentrations of any of the indicator
organisms monitored (E. coli, faecal streptococci, Candida albicans, dermatophytic fungi)
during a prospective epidemiological study at two beaches in Malaga, Spain. Heaney et al.
(2009) studied associations between sand exposure and illness at marine and freshwater beaches
under the U.S. EPA’s NEEAR study. The rates of illness varied from beach to beach, however,
overall the authors observed that digging in sand was associated with a modest increase in
reported gastrointestinal illness (adjusted incidence ratio 1.13; 95% confidence interval [CI]
1.02-125). Individuals who reported being buried in sand showed a slightly stronger incidence
of gastrointestinal illness (1.23; CI,1.05-1.43) and diarrhea (1.24; CI,1.01-1.52). No associations
were demonstrated for sand contact and non-enteric illness (Heaney et al., 2009).
Epidemiological studies in which microbiological sampling has been performed at very shallow
depths have similarly failed to show a correlation between water quality at this depth and
swimmer illness (Calderon et al., 1991; Fleisher et al., 1996; McBride et al., 1998; Haile et al.,
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1999). Microbiological counts in shallower waters can be expected to exceed those at greater
depths, owing to disturbances to foreshore and nearshore sands and sediments.
Managing health risks
Management actions to reduce the extent of faecal contamination affecting the beach area, as
well as steps to restrict swimmer exposure to recreational waters during periods of or in areas
perceived to be at increased risk are part of an effective strategy to protect against the risk of
human exposure to pathogens arising from faecal contamination in the foreshore and nearshore
sand environment.
The Environmental Health and Safety Survey is an important tool for helping recreational water
operators identify potential onshore sources of faecal contamination that are relevant to their
beach area. Further information on the EHSS process can be found in Part I (Management of
Recreational Waters).
For beach managers and operators, barriers to reduce the extent of faecal contamination can
include the physical removal of litter that may attract animals to the area and the installation of
physical barriers designed to discourage wildlife. Examples of such barriers can include animalproof refuse containers, fences and gull nets. Jurisdictional regulations restricting access for pets
on public beaches present another potential control mechanism.
Physical manipulations of the sand environment have also been proposed as a potential action to
help minimize faecal contamination and reduce its transport to swimming waters. Kinzelman et
al. (2004b) reported that deep mechanical grooming without levelling was effective in reducing
sand levels of E. coli at a Racine, Wisconsin, beach, particularly in wet sand. Targeted beach
grading to increase the steepness of the slope of a beach has been suggested as another action
that can improve water quality (City of Racine Health Department, 2006). A more steeply sloped
beach reduces the area vulnerable to wave swash and permits more rapid sand drying through
improved drainage (Clean Beaches Council, 2005; City of Racine Health Department, 2006).
Pre-emptive beach postings or swimming advisories restricting recreational water activities for
short periods immediately after rainfall events present another potential barrier. These act by
limiting swimmer exposure to faecal contamination that may have been washed from the sand
environment to the swimming area.
Beach users can also do their part to contribute to these strategies by ensuring that their litter is
properly disposed of, refraining from feeding animals on or near the beach, and complying with
any existing beach regulations or codes of conduct. They may also contribute by becoming
informed of steps that can be taken to reduce their personal exposure. Beach users are reminded
to adopt proper hygiene practices such as avoiding mouth contact with items that have been in
contact with sand, washing their hands prior to eating and showering as soon as is practical after
visits to the beach. The use of clean beach towels can also help reduce the degree of sand contact
(WHO, 2003a).
Larger-scale management options for beaches will require a comprehensive review of the
contamination inputs and watershed characteristics and the identification of specific options to
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minimize or control the sources of faecal contamination and to reduce the transfer of pollution to
the swimming area.
Summary
1. Beach sand and related environments may provide a more favourable environment for
microorganisms of faecal origin, which may permit them to survive for longer periods than in
the adjacent water. Physical factors such as wave action, storm surges, tidal activity and high
swimmer load can result in the transference of microorganisms from foreshore and nearshore
sand and sediments to swimming waters.
2. Currently, there is no conclusive evidence to indicate a link between microorganisms in
beach sand and illness among beach users. Further research is needed to determine the
relationships between faecal indicator bacteria and the possible presence of faecal pathogens
in beach sand, as well as the potential implications to human health.
3. Barriers that collectively reduce risk of exposure for beach users could include public
education campaigns, improved beach sanitation practices, appropriate sand grooming
practices and actions designed to discourage the activities of animals (birds and other
wildlife) within the beach area.
10.0 Faecal pollution source tracking
Description
A wide variety of faecal sources are capable of contaminating recreational waters. Faecal
pollution can come from sewage treatment plant effluents, stormwater and combined sewer
overflows, faulty or improperly designed septic tanks, improper farming practices, intensive
livestock and poultry operations, local wildlife (e.g., geese and gulls) and even recreational water
users themselves. Understanding the specific sources of faecal contamination can help to assess
public health risks and target appropriate risk management barriers. These, in turn, can reduce
beach postings and prevent potential waterborne disease outbreaks.
Faecal source tracking can be performed by a variety of chemical and microbiological methods.
Chemical methods
Numerous chemical compounds have been investigated as potential markers for human sources
of faecal pollution. Chemical methods of analysis are based on the detection of chemical
compounds known to be present in faecal material as a result of human activities—either through
consumption and/or metabolism and the subsequent excretion in faeces, or via disposal as
sewage wastes. Caffeine, detergents, laundry brighteners, fragrance materials, faecal sterols and
faecal stanols have been proposed as markers of faecal pollution from sewage treatment plants
(Glassmeyer et al., 2005). Chemical tracers such as dyes have also been used to confirm
suspected point sources of contamination, such as wastewater outfalls. A noted advantage of
using chemical markers is that the time required for analysis is shorter than that for many
microbial methods. There are still a number of information gaps regarding the use of chemical
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markers in terms of presence and fate in the environment and the sensitivity of the detection
methods. More research is currently needed to clarify some of the uncertainties surrounding their
use as potential faecal source indicators.
Microbial methods
Microbial source tracking (MST) is an emerging field. Methods are based on comparing the
similarity of microorganisms in water samples with those from known faecal sources to make
inferences about the source of faecal pollution. In recent years, a growing number of
microbiological methods have been developed (Scott et al., 2002; Simpson et al., 2002; Meays
et al., 2004; Edge and Schaefer, 2006). Consequently, more attention has been focused on
microbial methods than on chemical methods.
Microbial methods can be divided into library-dependent and library-independent methods.
To date, library-dependent methods have been more widely used in microbial source studies,
although library-independent methods are increasingly under investigation. Library-dependent
methods are based upon choosing a faecal indicator microorganism (e.g., E. coli or enterococci)
and establishing a reference library of characteristics of individual isolates obtained from known
faecal pollution sources. For example, a library could be a database of antibiotic resistance
profiles or DNA fingerprints of E. coli isolates obtained from animal faeces and municipal
wastewater effluent (Wiggins, 1996; Dombek et al., 2000; Carson et al., 2001). The similarity of
the profiles or fingerprints of E. coli isolates obtained from recreational waters (“unknowns”) can
then be compared with the profiles or fingerprints in the library (“knowns”) to make statistical
inferences about the source of the waterborne E. coli isolates.
Library-independent methods are based upon detecting host-specific markers to indicate the
presence of faecal contamination from a specific human or animal host in the water. Most
library-independent methods rely on PCR to detect the host-specific markers in water samples.
Host-specific markers may be such things as toxin genes (Khatib et al., 2002, 2003), genes for
virulence factors (Scott et al., 2005) or highly conserved DNA sequences (Bernhard and Field,
2000a). Some of the most promising results to date for developing host-specific markers for
faecal pollution source tracking involve 16S rDNA markers within the genus Bacteroides. These
anaerobic bacteria comprise a very large portion of the faecal flora in warm-blooded animals.
Bernhard and Field (2000b) developed Bacteroides 16S rDNA PCR assays specific to ruminants
and humans, and such assays have been applied successfully to MST studies in recreational
waters (Boehm et al., 2003; Bower et al., 2005; Noble et al., 2006).
An important aspect of MST has been discriminating between human and animal faecal
contamination. This is because faecal pollution from a human source (e.g., sewage) may present
human health risks different from those associated with faecal pollution from animal sources.
Viruses that infect humans are more likely to occur in human faecal wastes. However, wildlife
species can carry pathogens such as Campylobacter, Cryptosporidium and Giardia, which can
pose a risk to human health.
Studies in the late 1990s had raised considerable expectations for microbial source tools to
resolve problems in faecal pollution source identification. Recent publications (Griffith et al.,
2003; Stoeckel et al., 2004) have pointed out some of the limitations associated with the use of
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microbial methods. Library-based methods require very large reference libraries and can have
high misclassification rates. Non-library-based methods currently suffer from gaps in the
information on the host specificity of the markers. Additional research is required to understand
more fully the advantages and limitations of MST methods.
Current state of the science
Current thinking is that while many methods exist in the current MST toolbox, there is no
universally accepted best method. While some methods have achieved a level of maturity where
they could be considered for standardization, others are still experimental or research-grade tools
(Edge and Schaefer, 2006). It is further felt that with any MST study, it is advisable to have
multiple lines of evidence before making inferences about sources of faecal contamination.
Some recent MST studies have been perceived to have placed a “wet blanket” on the field by
illustrating some of the significant challenges associated with these types of undertakings. These
include the high costs of analysis and the difficulties encountered when using these technologies
in watersheds that have numerous faecal sources. Nevertheless, there have been successful
applications of MST in field studies. MST methods have been used to identify unexpected faecal
pollution sources, to verify information from other lines of evidence, to resolve local beach
closure problems involving limited contamination sources and to break down large source
tracking problems into more manageable studies. Edge and Hill (2007) demonstrated the
application of two library dependent MST methods for determining that bird droppings, rather
than municipal wastewater, were the primary source of faecal pollution responsible for beach
postings in Hamilton Harbour.
The field of MST is still evolving, and the MST toolbox is getting better. Novel molecular tools
such as DNA microarrays (Hamelin et al., 2006; Soule et al., 2006) and protozoan genotyping
methods (Jiang et al., 2005; Ruecker et al. 2007) could lead to the identification of new hostspecific DNA markers. Other tools based on DNA markers for host cells in faeces may also
prove useful for faecal pollution source tracking in the future (Martellini et al., 2005).
Application of microbial source tracking
MST studies can be expensive and time-consuming. As well, the current state of the science is
such that the methods may not be able to completely resolve all of the sources contributing to the
faecal contamination of the watershed and the recreational water area.
A good understanding and formulation of the nature of the faecal contamination problem are
required before considering the need for an MST study. The Environmental Health and Safety
Survey is a useful tool and an important first step for helping recreational water area operators,
service providers and local authorities identify the sources of contamination that are relevant to
their recreational water area.
At present, it is not possible to recommend a standard faecal source identification approach and
method that would be applicable to any situation in recreational waters. The decision of which
MST method to use will be influenced by factors such as the potential complexity and number of
faecal pollution sources, the geographic and temporal considerations for the study area, and the
available funds, equipment and expertise to conduct the study. For individuals wishing to pursue
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this approach, some guidance for the selection of an appropriate MST approach can be found in
publications from the U.S. EPA (2005b) and the U.S. Geological Survey (Stoeckel, 2005).
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Appendix A: List of acronyms and abbreviations
AK
ANC
bw
CDC
cfu
DAEC
DNA
EAEC
EHEC
EHSS
EIEC
EMPACT
EPA
EPEC
eq
ETEC
EU
GAE
GIS
HAV
HCGI
HUS
LD50
MAC
MPN
MST
NEEAR
NOAEL
NTU
PAM
PCR
pfu
QPCR
rDNA
RNA
S.A.F.E.
SRSV
STEC
TDI
UK
USGS
UV
VTEC
WHO
amoebic keratitis
acid-neutralizing capacity
body weight
Centers for Disease Control and Prevention (U.S.)
colony-forming unit
diffuse adherent E. coli
deoxyribonucleic acid
enteroaggregative E. coli
enterohaemorrhagic E. coli
Environmental Health and Safety Survey
enteroinvasive E. coli
Environmental Monitoring for Public Access and Community Tracking
Environmental Protection Agency (U.S.)
enteropathogenic E. coli
equivalents
enterotoxigenic E. coli
European Union
granulomatous amoebic encephalitis
geographic information system
hepatitis A virus
highly credible gastrointestinal illness
haemolytic uraemic syndrome
median lethal dose
maximum acceptable concentration
most probable number
microbial source tracking
National Epidemiologic and Environmental Assessment of Recreational
no-observed-adverse-effect level
nephelometric turbidity unit
primary amoebic meningitis
polymerase chain reaction
plaque-forming unit
quantitative polymerase chain reaction
ribosomal deoxyribonucleic acid
ribonucleic acid
Swimming Advisory Forecast Estimate
small, round, structured viruses
Shiga toxin-producing E. coli
tolerable daily intake
United Kingdom
United States Geological Survey
ultraviolet
verocytotoxigenic E. coli
World Health Organization
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Appendix B: Microbiological sampling and analysis
B.1
Sampling procedures for water
Samples for microbiological examination should be collected in clean, sterilized,
environmentally sensitive bottles with screw cap closures. Borosilicate glass or autoclavable
plastic bottles capable of withstanding repeated sterilization at 121°C or 170°C are
recommended. Bottles capable of holding volumes of 200–500 mL should be adequate for most
analyses; however, certain circumstances may require the collection of greater volumes of water
(e.g., 1 L, 10 L).
When sampling by hand, the sterilized bottle should be opened with the base firmly held in one
hand and with the opening facing downward. Exceptional care should be taken at all times
during sampling to avoid accidental contact with the mouth of the bottle or the bottle cap. The
bottle mouth is plunged downward into the water 15–30 cm below the surface (for both deep and
shallow waters). The bottle is then turned towards the current (if there is one), tilted slightly
upwards to displace the air and then gently pushed forward away from the hand, body, boat or
other sampling platform. Samples collected from a boat or other platforms should be collected
from the upstream side of these objects.
When collecting is done with a sampling pole, the bottle should be fit into the holder in the
recommended manner. With the cover removed, the sample should be collected upstream and
away from the collector by simulating the scooping method used during the collection of a handcollected sample.
The volume collected should be sufficient to enable all of the required testing to be performed.
Before recapping the bottle, a small amount of sample should be discarded, leaving an airspace
to allow for proper mixing prior to analysis. Once capped, the bottle should be properly labelled
and placed in an insulated cooler that contains pre-frozen cooling packs or ice. Additional data,
such as the time of sample collection, temperature of the water and similar observations, should
be recorded at this time as part of the sample record.
Responsible authorities may also wish to include within their monitoring program a requirement
to collect supplemental data on various water quality and meteorological parameters at the areas
being monitored. Several researchers have reported on the usefulness of such information in
developing mathematical models for predicting recreational water quality (Nevers and Whitman,
2005; Olyphant and Whitman, 2005). Potentially useful measurements may include the
following:
• amount of rainfall;
• degree of sunlight and cloud cover;
• temperature (air, water);
• tidal stage and water level;
• wave height;
• number of swimmers;
• wind direction and speed;
• bird populations (gulls, ducks, geese); and
• turbidity.
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An example of a sample collection and reporting form is provided in Appendix E.
B.2
Sampling procedures for sand and sediments
When epidemiological or other evidence indicates that swimming beaches could be the source of
waterborne diseases among swimmers, sand and sediment sampling and analysis for suspected
pathogens may be warranted. Many investigations have demonstrated that faecal indicator
bacteria and faecal pathogens can persist for extended periods in sand and sediments.
At present, there is no single preferred procedure for the collection and analysis of sand and
sediment samples. A variety of methods have been proposed for the collection of sand samples,
including the use of sterilized scoops, spatulas, corers, probes and other sterilized collection
vessels. Authorities are advised to consult the scientific literature to determine which methods
may be the most suitable for their needs. Sediment samples may be collected with sterile, 250- to
500-mL wide-mouth jars, and the same precautions that were used with water sampling should
be followed to ensure aseptic collection. In shallow waters, the jars are pushed along the bottom,
collecting the material at the sediment–water interface. When it has been half filled, the container
is retrieved, the excess water is poured off and the container is sealed. In deeper waters, sediment
samplers used for collecting benthic invertebrates, such as the Ponar, Petersen or Ekman grabs,
can also be used (APHA et al., 2005). Regardless of the method used, the use of sterilized
equipment and aseptic technique is of utmost importance in minimizing the risk of accidental
contamination of the sample. As with water samples, once collected, sand and sediment samples
should be properly labelled and placed in an insulated cooler containing pre-frozen cooler packs
or ice.
B.3
Sample transport, preservation and storage
After collection, samples should be held at temperatures below 10°C and in the absence of light
until the time of analysis. Insulated coolers containing pre-frozen cooling packs or ice may be
used during sample transport to the laboratory. To prevent the possibility of contamination,
samples should be packed in such a manner so as to prevent contact between the bottle lids and
any free water (from thawing/melting of the cooling material or otherwise) that may be retained
within the cooler. As well, the samples should never be frozen. Thus, they should be protected
from direct contact with the ice or cooling packs as necessary to prevent freezing.
Storing samples under refrigerated temperatures will have only a limited effect in terms of
preserving the distribution of microbiological populations in a recreational water sample. As a
result, microbiological analysis of water samples should be initiated as soon as possible to avoid
unpredictable changes in the microbial population (APHA et al., 2005). Samples should be
analysed within 24 hours from the time of collection (Bartram and Rees, 2000); however,
analysis within 8 hours is recognized as the preferred time interval (Bartram and Rees, 2000;
APHA et al., 2005). If the time required for transport is expected to exceed 6 hours, it is
recommended that field analysis be considered (APHA et al., 2005). Similarly, if the results are
to be used in legal action, it is recommended that special means (e.g., rapid transport, courier
service) be used to deliver the samples within the specified time limits and to maintain the chain
of custody (APHA et al., 2005).
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Data such as the temperature of the samples when received at the laboratory and the times of
sample collection, reception and analysis should be recorded as part of the sample record. Such
information may prove valuable during the interpretation of results (Bartram and Rees, 2000).
B.4
Methods for microbiological analysis
B.4.1 Recommended indicators of faecal contamination
Consideration should be given to the type of water being analysed when selecting the most
appropriate method for analysis. Currently, two main types of methods are used for the routine
detection and enumeration of E. coli and enterococci in recreational waters: the multiple-tube
fermentation (MTF) method and the membrane filtration (MF) method.
Multiple-tube fermentation (MTF) procedure
In this method, serial dilutions of a 100-mL sample are prepared in replicate sets of tubes or
wells containing differential media. These are then incubated and examined for the number of
positive test reactions. The number of positive tubes per dilution is then matched to a most
probable number (MPN) table, which provides a statistical estimate of the number of target
organisms in the original 100-mL sample.
An advantage of the MTF method is that it can be useful for samples where conditions have
made MF unusable—for example, with turbid, coloured or grossly contaminated water (Health
Canada, 2006a). As well, the use of liquid media may more easily permit the recovery of stressed
organisms. Limitations are that the method can be time-consuming, can require large amounts of
media and glassware and can require a long turnaround time to obtain results, particularly if
confirmation steps are required. As well, the method provides only a statistical estimate of the
presence of the target organism and not a true count of the total number of bacteria present.
Membrane filtration (MF) procedure
In this procedure, the water sample (usually 100 mL) is passed through a filter that retains the
bacteria. The filter is then placed on an appropriate differential or selective medium and
incubated. Following incubation, the appropriate colonies are counted, and the result is recorded
as the number of target organisms per 100 mL.
Advantages of the MF method are the ability to test greater volumes of water, reduced
requirements for labour and materials and the greater reliability and reproducibility associated
with a direct count. For these reasons, the MF procedure is largely favoured over the MTF
procedure for the routine examination of recreational waters. There are some limitations
associated with the MF method, however. Samples that are high in turbidity can interfere with
filtration, and highly contaminated samples may overwhelm the filter, preventing its ability to
produce an accurate count. As well, the direct transfer onto solid, selective media may hinder the
recovery of some organisms, and thus a resuscitation step may be required.
Defined substrate technologies
With both procedures, more recent methods have been based on the ability to detect specific
enzymes considered to be characteristic for the target organisms. E. coli methods are based on
the detection of the enzyme β-glucuronidase, which is thought to be restricted to this organism
and a few strains of Salmonella and Shigella. Enterococci methods are based on the detection of
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the enzyme β-glucosidase, which is characteristic for this group. Specifically tailored
chromogenic or fluorogenic substrates are incorporated into the growth media, which, when
metabolized by the target organism, confer a unique property to the developing colony or
surrounding media that can be used for diagnostic purposes. Chromogenic substrates produce a
distinct colour change when hydrolysed, whereas fluorogenic substrates produce a fluorescent
product that can be detected upon viewing under long-wavelength ultraviolet (UV) light.
Many currently available commercialized methods have used these principles in what has been
termed “defined-substrate technology.” With this technology, the indicator substrates are
specifically designed to function as both the principal source of carbon and energy for the target
bacteria. Other competitive bacteria cannot utilize the substrate and are therefore unable to
interfere with the recovery of the target.
Escherichia coli
Standard Methods for the Examination of Water and Wastewater (APHA et al., 2005)
recommends the mTEC method—an MF method developed by the U.S. Environmental
Protection Agency (EPA)—for the detection and enumeration of E. coli at natural swimming
beaches. This method involves placement of the filter on selective (mTEC) media, a resuscitation
step at 35°C for 2 hours to rejuvenate stressed organisms, incubation at 45°C for 22 hours to
detect all thermotolerant coliforms and then transfer to a urea substrate medium to differentiate
urease-negative E. coli from other thermotolerant coliforms, which are mostly urease-positive.
The U.S. EPA has published a list of approved methods for enumerating E. coli in recreational
waters (U.S. EPA, 2006c). Method 1103.1 is the original mTEC method. Method 1603 is a
modified (single-step) mTEC method that uses a single medium containing a chromogenic
substrate (5-bromo-6-chloro-3-indolyl-β-D-glucuronide, or BCIG). Hydrolysis of the BCIG
substrate by E. coli imparts a red or magenta colour to the resulting colonies, differentiating them
from other thermotolerant coliforms. Method 1604 is another MF method that uses MI medium,
which contains both a fluorogen (4-methylumbelliferyl-β-D-galactopyranoside, or MUGal) and a
chromogen (indoxyl-β-D-glucuronide, or IBDG) for the simultaneous detection of both total
coliforms and E. coli, respectively.
Other methods using similar principles have been approved for the enumeration of E. coli in
water supplies (ISO, 1998; APHA et al., 2005; U.S. EPA, 2006b). Similarly, various commercial
methods – prepackaged miniaturized or simplified versions of traditional MF/MTF tests – have
also received approval (U.S. EPA, 2006b). Laboratories in Canada may wish to evaluate the
applicability of specific methods to recreational waters in their regions.
Enterococci
Standard Methods for the Examination of Water and Wastewater (APHA et al., 2005) contains
two official methods for the examination of enterococci at swimming beaches—an MTF
procedure and an MF procedure. The MTF procedure involves inoculation of a series of tubes of
azide dextrose broth, incubation at 35°C for 48 hours and confirmation of tubes showing growth
on a bile-esculin-type agar by the appearance of brownish-black colonies with brown halos. The
MF procedure represents the mE method, originally described by the U.S. EPA in 1985. Filters
are placed on mE agar, incubated at 41°C for 48 hours and then transferred to esculin-iron agar
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(EIA) medium to confirm esculin hydrolysis. Pink to red colonies with black or reddish-brown
precipitates are considered enterococci.
The U.S. EPA has similarly published a list of approved methods for the enumeration of
enterococci from recreational waters (U.S. EPA, 2006c). Method 1106.1 is the original
mE method. Method 1600 is a modified (single-step) method (mEI), which reduces the time
of analysis from 48 hours to 24 hours. The method uses a single medium that contains a
chromogenic substrate (indoxyl-β-D-glucoside). Hydrolysis of the substrate by enterococci
confers a blue halo to the resulting colonies, distinguishing them from other non-enterococci.
As with E. coli, other methods using similar principles have been approved for the enumeration
of enterococci in water supplies (ISO, 2000; APHA et al., 2005; U.S. EPA, 2006b). Various
commercial methods – prepackaged miniaturized or simplified versions of traditional MF/MTF
tests – have also received approval (U.S. EPA, 2006b). Laboratories in Canada may wish to
evaluate the applicability of specific methods to recreational waters in their regions.
B.4.2 Pathogenic microorganisms
The routine testing of recreational waters for the presence of pathogenic microorganisms
(bacteria, viruses, protozoa) is not recommended. However, certain circumstances may warrant
testing for the presence of specific organisms, such as during investigations of potential
waterborne disease outbreaks.
Bacterial pathogens
Standard Methods for the Examination of Water and Wastewater (APHA et al., 2005) describes
methods for the isolation, detection and identification of the bacterial pathogens of concern for
recreational waters: Campylobacter, E. coli O157:H7, Salmonella, Shigella, Aeromonas,
Legionella, Mycobacterium, Pseudomonas, Leptospira and Staphylococcus aureus.
In general, methods for the isolation and detection of bacterial pathogens in recreational waters
follow the same basic design: concentration of the target organism (through MF, centrifugation
or growth in enrichment media), differentiation from non-target organisms (e.g., via growth in
selective media or using antibody-based methods of detection) and confirmation (through a
combination of morphological assessment, response to biochemical tests and serological
identification).
Methods for more rapid detection of bacterial pathogens in recreational waters using advanced
biochemical, immunological or gene sequence-based technologies are currently being explored.
Notably, developments in polymerase chain reaction (PCR)-based methods have generated
interest in the application of this technology for such purposes. Currently, PCR methods have
been described for the waterborne detection of all of the bacterial pathogens considered to be of
concern. Similarly, quantitative or real-time PCR methods have been described for a number of
important enteric bacterial pathogens, including Salmonella, Campylobacter and E. coli
O157:H7. At present, more work is needed in this area to develop standardized methods that can
be accurately, reliably and affordably used. Several authors have published reviews describing
the current state of the knowledge with respect to emerging technologies for the enumeration and
detection of recreational water pathogens (Ashbolt, et al., 2001; Noble and Weisberg, 2006;
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Savichtcheva and Okabe, 2006). Laboratories in Canada are advised to consult the literature for
further information as to the potential applicability of specific methods to recreational waters in
their areas.
Viral pathogens
The detection of pathogenic viruses is beyond the scope of most water microbiology laboratories.
When necessary, testing should be performed only by trained virologists having adequate
facilities for the proper handling of these organisms.
The methodology for detection of waterborne viruses has been standardized to a certain extent;
however, even the most accepted methods continue to be researched, modified and improved.
Standard Methods for the Examination of Water and Wastewater (APHA et al., 2005) contains
procedures for the concentration of viruses and provides guidance on methods for virus detection
and identification. As well, the U.S. EPA has published a Manual of Methods for Virology
(U.S. EPA, 2001a), which provides detailed, step-by-step procedures for the recovery, detection,
enumeration and identification of viruses from water, sewage and other related effluents.
In general, the recovery and detection of pathogenic viruses from surface waters samples are
difficult processes. Viruses are present in small numbers in faecally contaminated waters; thus,
large volumes of water (up to thousands of litres) must be concentrated in order to detect their
presence. Adsorption-elution methods or ultrafiltration represent the main methods used to
collect and concentrate viral particles from water samples. Further concentration of the filtered
sample can be accomplished through the application of flocculation or precipitation techniques.
Traditionally, cell culture methods have been the most widely used for the detection of enteric
viruses in water. These methods provide valuable information on virus infectivity and
concentrations, yet they can be laborious and time-consuming, and not all viruses can be grown
in cell culture. More recently, conventional PCR-based methods have been applied to the
detection of viruses in environmental samples. Currently, no suitable alternative to cell culture
exists to enable the assessment of virus viability. Specific applications of PCR technology
(reverse transcriptase PCR, quantitative PCR, integrated cell culture PCR) are being explored for
their ability to overcome some of the limitations associated with conventional PCR methods.
Numerous variations of the PCR method have been described that have been successfully used
for the detection of viruses in recreational waters. Although much progress has been made in this
area, standardized procedures have not yet been developed. Individuals are thus advised to
consult the literature for further information regarding specific methods. Review articles
summarizing the existing methods as well as potentially emerging technologies have been
published (Bosch, 1998; Griffin, et al., 2003; Fong and Lipp, 2005).
Protozoan pathogens
Testing for pathogenic protozoa is outside the scope of analytical services provided by most
water testing laboratories. Analysis requires the use of trained analysts and highly specialized
equipment.
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Standard Methods for the Examination of Water and Wastewater (APHA et al., 2005) does
provide an overview of the methods that can be used for the recovery and detection of Giardia
and Cryptosporidium in environmental samples; however, specific procedures are not described.
The U.S. EPA has approved two methods for the detection of Giardia and Cryptosporidium in
water samples: Method 1622, which is a stand-alone method for Cryptosporidium, and Method
1623, which can be used for the simultaneous detection of both organisms (U.S. EPA, 2006c).
These methods are considered the most widely used for the detection of Giardia and
Cryptosporidium in water. The detection of Giardia cysts and Cryptosporidium oocysts is a
complex and difficult procedure. Even the most widely used and accepted methods suffer from
limitations associated with the recovery of these organisms. Work is ongoing to continue to
refine and improve the accuracy and sensitivity of these methods.
The recovery and detection of Giardia cysts and Cryptosporidium oocysts in water samples
comprise three basic steps: concentration (via flocculation, centrifugation or filtration),
separation from interfering debris (through the use of density-gradient centrifugation,
immunomagnetic separation or fluorescence activated cell sorting) and detection (via
immunofluorescent staining or PCR-based methods).
Of the methods used for detection, immunofluorescent staining currently represents the most
widely used technique. This procedure involves the application of fluorescent antibodies directed
against cyst and oocyst antigens, followed by identification of the labelled cysts and oocysts
under an immunofluorescence microscope.
Various conventional PCR-based methods have been described for the detection of Giardia and
Cryptosporidium in recreational waters. Other analytical techniques, such as restriction fragment
length polymorphism (RFLP) analysis, have been described for the further characterization of
Cryptosporidium species and genotypes. Genotype information can be used to help identify the
potential host sources of Cryptosporidium responsible for an outbreak.
One limitation of the current detection methods is that they do not provide information on the
viability or infectivity of the cysts or oocysts. Separate assays have been described for these
purposes, which involve observing the degree of excystation or the inclusion/exclusion of
specific fluorescent dyes. Other methods require the use of cell culture or animal subjects.
Specific variants of the PCR method (reverse transcriptase PCR, quantitative PCR) have also
been designed to facilitate quantification of the organisms and to provide estimates of cyst or
oocyst viability. In general, these tests can be expensive and difficult to perform and are typically
reserved for specific research purposes.
Readers are advised to consult the literature for further information regarding specific methods.
Published reviews of molecular techniques used for the detection and identification of Giardia
and Cryptosporidium are similarly available (Caccio, 2003).
Cyanobacteria and their toxins
Several methods are available for the detection of cyanobacteria and microcystins in water and
bloom samples. These methods vary markedly in their level of complexity and the level of
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Guidelines for Canadian Recreational Water Quality: Third Edition
information that they provide. Selection of an appropriate method will depend on the type and
degree of information required as well as the availability of laboratory facilities and experienced
personnel.
Published texts are available that discuss the methods that may be used for the detection of
cyanobacteria and microcystins in recreational waters (Chorus and Bartram, 1999; Falconer,
2005). Review articles summarizing the existing and potentially emerging technologies have
similarly been published (e.g., McElhiney and Lawton, 2005). Individuals are advised to consult
the literature for information on specific methods.
Cell counting
Standard Methods for the Examination of Water and Wastewater (APHA et al., 2005) contains
procedures for the enumeration of cyanobacteria (as phytoplankton) in water. Concentrations are
determined by direct microscopic count with a counting chamber of known dimensions and then
back-calculating to the volume of the original sample. The majority of cyanobacteria are present
as colonies or filaments, which can be difficult to distinguish; thus, the use of a trained
microscopist with experience in identifying cyanobacteria is recommended.
Microcystin testing
Total toxin analysis requires the determination of both the free and cell-bound microcystins.
Thus, initial processing steps are often required to extract and concentrate the toxins in the
sample. These may include concentration of the cyanobacterial cells, cell lysis and toxin
extraction and purification. A field (commercialized) test kit could be used as a screening tool to
determine the presence or absence of toxin in a water supply. If the presence of toxin is detected
in a sample using the field test kit, a sample can then be submitted to a recognized laboratory for
confirmatory analysis.
The mouse bioassay had historically been used as the primary method for providing evidence of
a bloom’s toxicity. In recent years, it has been replaced by more sensitive and reliable laboratory
methods, such as the protein phosphatase inhibition assay and enzyme-linked immunosorbent
assay (ELISA) tests. One drawback of these methods is that they do not provide direct evidence
of the sample’s toxicity. A number of invertebrate and cell culture assays have been investigated
as possible alternatives for the mouse bioassay.
The protein phosphatase inhibition assay is a biochemical assay used for the detection of
microcystins based on their ability to inhibit the activity of protein phosphatase enzymes. It is
a rapid, sensitive method for the detection of microcystins. The test provides a quantitative
estimate of the total microcystins present, but does not provide information on the composition
of microcystin variants within the sample. A commercialized kit for the detection of microcystins
in water samples is currently available
The ELISA technique is regarded as a very promising method for the detection of microcystins
in bloom and water samples. ELISA methods are rapid and sensitive, but are presently not
specific enough to distinguish between individual microcystin variants. A number of
commercialized kits for the detection of microcystins in water samples are currently available.
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Guidelines for Canadian Recreational Water Quality: Third Edition
Liquid chromatography (LC) coupled with mass spectroscopy (MS) is the most commonly used
laboratory method for the identification and quantification of microcystins and represents the
reference standard against which other methods are judged. Analysis using this method is time
consuming, technically demanding and expensive as it requires specialized equipment.
However, standardized procedures have been described, and many analytical laboratories possess
the necessary instrumentation. The lack of available standards has limited this method’s
usefulness in identifying individual microcystin variants.
Recently, PCR-based methods have been described for the detection of cyanobacteria based
on primers directed against gene fragments belonging to Microcystis spp., as well as the genes
responsible for microcystin production (mcy genes). These methods have demonstrated some
measure of success in the identification and quantification of cyanobacteria and their toxins in
bloom samples; however, the techniques are still under development, and further investigation
is required.
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Guidelines for Canadian Recreational Water Quality: Third Edition
Appendix C: Composite sampling for faecal contamination
C.1
Description
The Guidelines recommend that when assessing water quality for adherence with the guideline
values for the recommended indicators of faecal contamination, a minimum of five samples
should be collected, at a minimum frequency of one sample per week.
Composite sampling (collection of multiple samples from across a stretch of beach, combining
the samples into one large composite, analysing a subsample of the resulting mixture) presents a
possible means for increasing the area covered under a beach monitoring program, while
potentially minimizing the costs associated with analysis.
Authorities and beach operators and/or service providers may wish to investigate composite
sampling as a tool to potentially improve the utility of their faecal indicator monitoring
programs.
C.2
Introduction
With some monitoring programs, single samples may be used to characterize long stretches of
swimming water, across several days of swimmer activity. It is known that the fluctuation of
water quality can be significant, even over short distances and time periods. Collecting multiple
samples more frequently is recommended, because increasing the number of samples increases
the reliability of the data (Whitman and Nevers, 2004). The costs associated with increased
monitoring, however, can be prohibitive. One solution that has been proposed to address this
problem is the use of composite sampling (U.S. EPA, 2005a).
The process of composite sampling requires collecting multiple samples. Equal volumes from
each sample are then mixed together to form a composite, which is then analysed as a single
sample. This technique can broaden the coverage of a sampling strategy where the analysis of
large numbers of samples would otherwise be required. Subsequently, it can be used to increase
sampling reliability without significantly affecting the costs associated with monitoring (Patil,
2002). Composite sampling has numerous applications in biomonitoring and environmental
sampling and has been used for assessing contamination in a variety of media, including soils,
air, water and biological tissue. Recently, investigations have been conducted to explore whether
composite sampling can be applied to the assessment of the quality of recreational waters
(Kinzelman et al., 2006). Preliminary evidence has indicated that when properly conducted,
composite sampling can be used in making water quality decisions with a degree of accuracy
comparable to that of traditional sampling regimens.
There are challenges associated with composite sampling that need to be taken into consideration
before introducing this technique into a recreational water monitoring program. These are
described briefly below.
Potential sources of bias
Composite sampling adds another layer of uncertainty to the water quality results, since a
subsample is being used to estimate the average indicator density over all samples, and this
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Guidelines for Canadian Recreational Water Quality: Third Edition
estimate in turn is used to characterize the water quality for the whole beach. It is suggested that
including more samples in the composite compensates for the effects of this bias.
There is the potential for the presence of an individual sample with a high concentration to be
masked when combined with samples with lower concentrations, owing to the effects of dilution
(Kinzelman et al., 2006). Samples from hot spots (areas where poor water quality is likely to be
persistent) should not be composited with other samples. The identification of hot spots may be
determined by conducting an Environmental Health and Safety Survey (EHSS) or through an
initial period of intensive sampling. In programs where hot spots have been characterized and
determined as unlikely to occur, the occurrence of a single sample with high concentrations may
be considered to be the result of natural, random variability (Kinzelman et al., 2006).
Comparing results with the guideline values
The composite sampling result approximates the arithmetic mean of the indicator counts of the
individual samples. However, when analysing bacteriological water quality data, the geometric
mean is recommended as the best estimate of central tendency of microbial populations. The
guideline values for the recommended indicators of faecal contamination are based on geometric
mean values. In determining whether the results are in accordance with the guidelines,
operators/service providers or responsible authorities would need to convert the composite result
to an approximation of the geometric mean.
Wymer et al. (U.S. EPA, 2005a) pointed out that the difference between the composite value and
the geometric mean can be compensated if the variance (v) of the log10 indicator densities is
known. An estimate of the variance can be calculated from historical log10 data. Once the
variance has been determined, multiplying the count per 100 mL obtained from the composite
sample by the factor 10−1.15v produces a value that is approximately equivalent to the geometric
mean of the individual samples (U.S. EPA, 2005a).
C.3
Study results
Kinzelman et al. (2006) produced swimming beach water quality data comparing the accuracy of
composite sampling with traditional monitoring practices at two Lake Michigan beaches in
Racine, Wisconsin. Water samples were collected over 68 days in 2003 from two public
swimming beaches and analysed for E. coli using single-sample analysis with arithmetic and
geometric mean averaging and composite sample analysis.
The resulting data indicated that composite sampling appeared to be an effective alternative to
traditional monitoring procedures. In general, the value of the composite sample fell within the
range of the single-sample values, and the data indicated an approximate 1:1 ratio between the
composite sample and the arithmetic mean of the individual samples. In comparing what would
have been the ultimate management decision (i.e., issue a water quality advisory or allow the
beach to remain open) resulting from the use of composite sampling versus the results from
individual analyses (singly or with averaging), the outcome for both methods remained constant
at one beach and differed in only two instances at the other. Compositing resulted in additional
advisories in both instances. Therefore, compositing appeared to introduce neither bias nor
additional variability into the monitoring results (Kinzelman et al., 2006). Verification studies,
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Guidelines for Canadian Recreational Water Quality: Third Edition
performed on a smaller scale in subsequent years, have yielded similar results, and Racine has
successfully used composite sample analysis for compliance monitoring since 2004.
C.4
Conclusions
Under the appropriate circumstances, compositing of samples may present a viable alternative to
current monitoring schemes that employ a single sample to characterize water quality over long
stretches of swimming beaches. Composite sampling may encourage more sampling to take
place, thus expanding the coverage of the monitoring program and increasing sampling
reliability, while at the same time maintaining the costs associated with monitoring or even
lowering them. Monitoring programs that require a large number of samples to be analysed could
benefit from adoption of this approach.
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Guidelines for Canadian Recreational Water Quality: Third Edition
Appendix D: Recreational swimming area Environmental Health and Safety
Survey (EHSS) checklist
Identification
Beach Name: __________________________________________________________
Address:
__________________________________________________________
__________________________________________________________
Responsible
Authority:
__________________________________________________________
__________________________________________________________
Tel.: ____________ Fax: ____________ E-mail: __________________
Person(s) Conducting
Survey:
__________________________________________________________
Date:
_____________________
Time:
_______________________
Background Information
Water Body Type:
____________________________________________
Dimensions of Beach:
Dimensions of Swimming
Area:
Number of Sampling Sites:
Length (m): _________
Length (m): _________
Width (m): __________
Width (m): __________
________
[Attach Map or Aerial Photo of Suitable Scale (including location of sample sites)]
Water Temperature
Prevailing Winds
Prevailing Currents
Seasonal Rainfall
Wave Height
High/Low (°C): ________
Direction:
________
Direction:
________
Total (mm):
________
Average (m): ________
Surrounding Land Uses (check all that apply):
Urban
□ Rural
□
Residential
□ Forest
□
Field
□ Hills/Uplands
□
Marsh/Swamp
□ Landfill
□
Harbour
□ Other: _________ □
Average (°C):
Avg. Speed (km/h):
Avg. Speed (km/h):
24-h High (mm):
Range (m):
_______
_______
_______
_______
_______
Agricultural (specify): ____________
Commercial (specify): ____________
Industrial (specify): ______________
River/Stream/Ditch:
Other: ________________________
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□
□
□
□
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Guidelines for Canadian Recreational Water Quality: Third Edition
Microbiological Hazards
Potential Sources of Faecal Contamination
Municipal Sewage Discharges
Stormwater Drains/Discharges
Wastes from Animal Feeding
Operations
Other Discharges Containing Faecal
Wastes (List):
_____________________
_____________________
□
□
□
□
□
Combined Sewer Overflows (CSOs) □
Septic Waste Systems
□
Other Sewage Collection/Disposal/
Treatment Systems (List):
_____________________
_____________________
□
□
Areas Receiving Sewage Sludge
Other: ___________________
Other: ___________________
□
□
□
Stormwater Runoff from:
Agricultural Areas
Beach and Surrounding
Facilities (e.g., parking)
□
□
Other Environmental Sources:
Discharging Rivers/Streams/Creeks □
Birds (e.g., gulls, ducks, geese, other) □
Other wild animals
□
Pets
□
Swimmers
□
Other: ___________________
□
Other: ___________________
□
(#’s: None Low Med High [circle one])
(#’s: None Low Med High [circle one])
(#’s: None Low Med High [circle one])
(#’s: Low Med High [circle one])
Items for Consideration during the Resulting Risk Assessment:
•
•
•
•
•
•
•
•
•
Proximity of potential contamination sources to the swimming area.
Potential for contamination sources to have an impact on the swimming area
(including an indication of their risk priority: Low, Medium, High).
Evaluation of water quality according to historical microbiological data (e.g.,
frequency of exceedances of the guideline values for the recommended indicators of
faecal contamination [e.g., continuous/periodic/sporadic]).
Discharges: Assessment of such factors as volume, flow rate, treatment type,
applicable indicator standards, periodicity (continuous, sporadic) and predictability.
Effects of rainfall: Levels triggering contamination events and typical event duration.
Assessment of swimming area circulation: Effect of onshore winds, tides, currents,
flow patterns in transporting faecal contamination to and entrapping it within the
swimming area.
Animals and birds: Assessment of their types, numbers and droppings.
Impact of swimmers on water quality—numbers, ages.
Assessment of potential barriers: Barrier types and points at which they may be
applied to reduce impact of the contamination source and/or swimmer exposure.
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Guidelines for Canadian Recreational Water Quality: Third Edition
Chemical Hazards
Potential Sources of Chemical Contamination
Commercial/Industrial Discharges
Motorized Watercraft
Other: _____________________
□
□
□
Marinas
Other: _____________________
□
□
□
Urban Areas
Other: ___________________
□
□
Stormwater Runoff From:
Areas subject to Pesticide
Application
Areas subject to Fertilizer
Application
□
Items for Consideration during the Resulting Risk Assessment:
•
•
•
•
•
•
•
Proximity of potential contamination sources to the bathing area.
Potential for contamination sources to have an impact on the swimming area
(including an indication of their risk priority: Low, Medium, High).
Discharges: Assessment of such factors as volume, flow rate, treatment type,
periodicity (continuous, sporadic) and predictability.
Effects of rainfall: Levels triggering contamination events and typical event duration.
Assessment of swimming area circulation: Effect of onshore winds, tides, currents,
flow patterns in potentially transporting chemical contamination to and entrapping it
within the swimming area.
Motorized watercraft: Assessment of their types and numbers.
Assessment of potential barriers: Barrier types and points at which they may be
applied to reduce impact of the contamination source and/or swimmer exposure.
Other Biological Hazards
Other Biological Hazards Known to Affect the Recreational Water Area
(Presence may be continuous, seasonal or sporadic.)
Cyanobacterial Blooms
Large Numbers of Aquatic Plants
Other (specify): _____________
□
□
□
Schistosomes (Swimmer’s Itch)
Other (specify): _____________
□
□
Items for Consideration during the Resulting Risk Assessment:
•
•
•
Seasonal nature of the hazard: continuous, annual, sporadic.
Presence of contributing factors (as applicable): water conditions, local geography,
temperatures, nutrient levels, presence of appropriate host species.
Assessment of potential barriers to control hazard and/or reduce human exposure in
areas/during times of increased risk.
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Guidelines for Canadian Recreational Water Quality: Third Edition
Physical Hazards and Aesthetic Considerations
Subsurface Hazards:
Steep Slopes or Dropoffs
Large Rocks
Other: _____________________
Other: _____________________
□
□
□
□
Depths greater than 4.5 m
Slippery or Uneven Bottom
□
□
Water Conditions:
Strong Currents or Rip Tides
□
Undertows
□
Other:
Litter on Beach
Floating Debris
Broken Glass or Other Sharp Objects
Medical Wastes
Seaweed/Algae on Beach
□
□
□
□
□
( None Low Med High [circle one])
( None Low Med High [circle one])
( None Low Med High [circle one])
( None Low Med High [circle one])
( None Low Med High [circle one])
Vehicles Permitted on Beach or Near Bathing Area:
Automobiles
Y/N
Boats/Watercraft
Y/N
Specify:___________________________
Items for Consideration during the Resulting Risk Assessment:
•
•
•
•
•
•
Assessment of the physical characteristics of the beach and their potential impacts on
safe enjoyable use of the area. Includes evaluation of physical layout (geography,
topography), composition of shoreline and bottom material, influence of existing
structures.
Assessment of potential risks posed by specific hazards/factors in causing injury or
illness or otherwise interfering with the enjoyable use of the area.
Shoreline and water free from obstructions and of sufficient clarity to permit viewing
of persons who may be in distress.
Assessment of the nature and origin of litter and floating debris.
Applicable physical and aesthetic parameters (pH, temperature, turbidity, colour,
clarity, litter) in agreement with recommendations given in the Guidelines for
Canadian Recreational Water Quality.
Assessment of potential barriers to control hazard and/or reduce human exposure in
areas/during times of increased risk.
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Guidelines for Canadian Recreational Water Quality: Third Edition
Facilities and Safety Provisions
Facilities:
Toilets
#: ____ □
Drinking Water Fountains #: ____ □
Other:
#: ____ □
Showers
Litter Bins
Other
#: ____ □
#: ____ □
#: ____ □
Lifesaving Equipment
First Aid Stations
#: ____ □
#: ____ □
Access for Persons with Disabilities □
Safety Provisions:
Lifeguard Stations
Emergency Telephone
#: ____ □
#: ____ □
Signs/Communication Materials:
Beach Posting/Suitability for
Swimming
Other Hazards (list):
_____________________
_____________________
□
□
□
Emergency Contact Information
□
Other: _____________________
□
Formal Procedures or Reporting Mechanisms in Place to Deal with:
Municipal or Industrial Spills/Discharges/Treatment Bypasses
Waterborne Disease Outbreaks
Swimmer Injuries
Items for Consideration during the Resulting Risk Assessment:
•
•
Assessment of the adequacy of facilities and safety provisions.
Evaluation of signs and other materials for public communication: Message clear
and concise, signs placed in locations highly visible to the public.
152
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□
□
Date:Edition
Guidelines for Canadian Recreational Water Quality: Third
Time:
Appendix E: Example: Sample collection and reporting form
Beach Name:
_________________________
Address:
_____________________________
_____________________________
Responsible
Authority:
_________________________
_________________________
Contact
_____________________________
Information: _____________________________
_____________________________
Person Collecting
Sample:
_________________________
Site
Indicator Counts
□ E. coli
□ Enterococci
Wind Direction: □ None
□ Onshore
Sunlight:
□ Offshore
□ Parallel to Shore
□ Sunny
□ Overcast
□ Partially Cloudy □ Rainy
Swimmer Density: □ None
□ Low
□ Medium
□ High
Approximate Number: ______
Geometric Mean:
Period Covered:
Air Temperature (°C):
___________________
Water Temperature (°C):
___________________
pH:
___________________
Salinity:
___________________
Turbidity:
□ Low
□ Medium
□ High
Value (NTU): _____________
Wave Activity:
□ Low
□ Medium
□ High
Height (m): ______________
Rainfall:
(during
sampling)
□ None
□ Low
Rainfall:
(Past 48 h)
□ None
□ Medium
□ Low
□ High
Value (mm): ______________
Birds:
▪ gulls
▪ ducks
▪ geese
□ None
Litter:
□ None
□ Low
□ Medium
□ High
Seaweed/
□ None
Algae (on beach)
□ Low
□ Medium
□ High
Notes:
□ Medium
□ High
153
□ Low
□ Medium
□ High
Approximate Number: ______
Guidelines for Canadian Recreational Water Quality: Third Edition
Appendix F: Examples: Informative beach signs
F.1
Beach posting: example of an informative swimming advisory sign
WARNING
WATER NOT SUITABLE
FOR SWIMMING
Clear
communication
that a risk
exists.
Instructions for
users regarding
the suitability
of the water.
Accompanying
graphic that is
easy to
understand.
Statement
regarding the
reason for the
posting.
FAECAL INDICATOR BACTERIA LEVELS EXCEED
RECOMMENDED LIMITS
CONTACT WITH THESE WATERS MAY CAUSE
ILLNESS
FOR MORE INFORMATION ON THE MONITORING
OF THIS BEACH, PLEASE CONTACT:
[Department – Telephone No.]
[NAME OF ISSUING AUTHORITY]
[CONTACT INFORMATION]
154
Additional
information
beneficial to
the public’s
understanding
of the situation.
Contact for more
information on
the details of the
monitoring
program.
Contact
information
for authority
responsible
for issuing
posting.
Guidelines for Canadian Recreational Water Quality: Third Edition
F.2
Beach posting: example of an informative water suitable for swimming sign
WATER OPEN
SUITABLE FOR
SWIMMING
Clear
communication
that the beach is
open.
Instructions for
users regarding
the suitability
of the water.
Accompanying
graphic that is
easy to
understand.
Statement
regarding the
reason for the
decision.
THE MOST RECENTLY COLLECTED SAMPLES INDICATED
THAT FAECAL INDICATOR BACTERIA LEVELS WERE BELOW
THE RECOMMENDED LIMITS
FOR MORE INFORMATION ON THE MONITORING OF THIS
BEACH, PLEASE CONTACT:
[Department – Telephone No.]
[NAME OF ISSUING AUTHORITY]
[CONTACT INFORMATION]
155
Contact for more
information on
the details of the
monitoring
program.
Contact
information
for authority
responsible
for issuing
posting.
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