Microbiological risk assessment: a scientific basis for managing drinking water safety from source to tap
Pathogens in drinking water sources
November 2004
Pathogens in drinking water sources
Authors:
Kathy Pond, Joerg Rueedi, Steve Pedley
Author affiliation:
Robens Centre for Public and Environmental Health
University of Surrey
Guildford, Surrey
United Kingdom
This study has been performed as part of the
MicroRisk project that is co-funded by the
European Commission under the Fifth Framework
Programme, Theme 4: “Energy, environment and
sustainable development” (contract EVK1-CT2002-00123).
The authors are solely responsible for the content
of this document. This document does not
necessarily represent the opinion of the European
Community and the European Community is not
responsible for the use of the information
appearing in this report.
2
Table of contents
Table of contents ...........................................................................................................3
List of Figures ................................................................................................................4
List of Tables ..................................................................................................................4
1
Background...........................................................................................................6
1.1
1.2
Waterborne diseases .........................................................................................................6
Water supply in Europe ......................................................................................................7
2
Analytical Methods ...............................................................................................9
2.1
2.2
Methodological considerations .........................................................................................13
Conclusions .....................................................................................................................17
3
Review of basic knowledge of the sources and occurrence of chosen
pathogens............................................................................................................18
3.1
3.2
3.2.1
3.2.2
3.2.3
3.2.4
3.2.5
3.2.6
3.2.7
3.2.8
3.3
3.3.1
3.3.2
3.3.3
3.3.4
3.4
Reservoirs of chosen pathogens ......................................................................................18
Potential Sources of contamination ..................................................................................18
Overview .............................................................................................................................................18
Hosts of pathogens and health implications .......................................................................................19
Cryptosporidium..................................................................................................................................20
Giardia ................................................................................................................................................22
Campylobacter ....................................................................................................................................22
E. coli 0157:H7 ...................................................................................................................................23
Enteroviruses .......................................................................................................................................24
Norovirus.............................................................................................................................................25
Pathogen loads in sewage and manure ...........................................................................25
Overview .............................................................................................................................................25
Cryptosporidium and Giardia.............................................................................................................26
Campylobacter and E.coli O157.........................................................................................................27
Enterovirus and Norovirus..................................................................................................................28
Summary..........................................................................................................................28
4
Persistence of pathogens in the environment.................................................29
4.1
4.1.1
4.1.2
4.1.3
4.1.4
4.1.5
4.1.6
4.1.7
Persistence of pathogens in surface waters .....................................................................29
Temperature.........................................................................................................................................29
Salinity.................................................................................................................................................32
Pressure................................................................................................................................................32
pH 32
Solar radiation and inactivation of pathogens ....................................................................................33
Ammonia .............................................................................................................................................33
Predation of pathogens........................................................................................................................34
5
Review of knowledge on transport of pathogens............................................36
5.1
5.1.1
5.1.2
5.2
5.2.1
5.2.2
5.2.3
5.2.4
5.2.5
5.2.6
Transport in surface water................................................................................................36
Settling.................................................................................................................................................36
Transport in sediments ........................................................................................................................37
Transport in the subsurface..............................................................................................37
Introduction .........................................................................................................................................37
Time scales of groundwater transport.................................................................................................38
Transport mechanisms ........................................................................................................................40
Modelling approaches .........................................................................................................................49
Transport through the unsaturated soil ...............................................................................................53
Transport through the saturated zone .................................................................................................55
3
6
Review of public domain information on contamination level of all types of
source water in the European Union ................................................................56
6.1
6.2
6.3
6.4
6.5
6.6
Cryptosporidium ...............................................................................................................56
Giardia .............................................................................................................................59
Campylobacter .................................................................................................................59
E. coli 0157 ......................................................................................................................61
Enteroviruses ...................................................................................................................61
Norovirus..........................................................................................................................63
7
Conclusions ........................................................................................................65
8
References ..........................................................................................................67
List of Figures
FIGURE 1 INACTIVATION RATES OF E. COLI AS FUNCTION OF TEMPERATURE. ....................31
FIGURE 2 INACTIVATION RATES DIFFERENT ENTEROVIRUSES ..................................................32
FIGURE 3 WATER RESIDENCE TIME IN INLAND FRESHWATER BODIES (AFTER MEYBECK
ET AL. 1989) .......................................................................................................................................38
FIGURE 4 ROCK TEXTURE AND POROSITY OF TYPICAL AQUIFER MATERIALS (BASED ON
TODD, 1980). A) WELL-SORTED, UNCONSOLIDATED SEDIMENT WITH HIGH
POROSITY (E.G. ALLUVIAL SANDS); B) POORLY SORTED SEDIMENT WITH LOW
POROSITY; C) WELL-SORTED SEDIMENT OF POROUS PEBBLES; D) SEDIMENT WHOSE
POROSITY HAS BEEN DIMINISHED BY DEPOSITION OF MINERAL MATTER; E) ROCK
WITH POROSITY INCREASED BY SOLUTION (E.G. LIMESTONE); AND F) ROCK WITH
POROSITY INCREASED BY FRACTURING (E.G. GRANITE)..................................................39
FIGURE 5 RANGE OF HYDRAULIC CONDUCTIVITY VALUES FOR GEOLOGICAL MATERIALS
(BASED ON DRISCOLL, 1986 AND TODD, 1980) .......................................................................41
FIGURE 6 DISPERSION IN A HOMOGENEOUS ISOTROPIC AQUIFER. A FIXED VOLUME OF
TRACER IS RELEASED AT THE INJECTION POINT A AT TIME 0. AT TIME T THE
TRACER HAS REACHED B; AFTER TIME T’ IT HAS REACHED C; AFTER TIME T’’ IT
HAS REACHED D (AFTER PRICE, 1996)......................................................................................42
FIGURE 7 PATHOGEN DIAMETERS COMPARED TO AQUIFER MATRIX DIAMETERS..............43
FIGURE 8 1-DIMENSIONAL ANALYTICAL SOLUTION OF TRACER CONCENTRATION AT
X=6M AWAY FROM THE INJECTION POINT, AVERAGE GROUNDWATER VELOCITY
OF 1M/DAY AND A DISPERSION OF 2M2/DAY. THE FULL CURVE INDICATES THE
SOLUTION FOR L=0DAY-1 AND THE DOTTED CURVE SHOWS THE RESULT FOR
L=0.5DAY-1. THE VERTICAL LINES SHOW THE TIME OF PEAK ARRIVAL AT THE
OBSERVATION POINT....................................................................................................................50
List of Tables
TABLE 1 WATERBORNE PATHOGENS AND THEIR SIGNIFICANCE IN WATER SUPPLIES.
SOURCE: WHO (2004) ..........................................................................................................................6
TABLE 2 PROPORTION OF GROUNDWATER IN DRINKING WATER SUPPLIES IN SELECTED
EUROPEAN COUNTRIES (EEA, 1999; UN ECE, 1999)....................................................................7
TABLE 3 METHODS FOR THE DETECTION OF MICROBIAL CONTAMINATION IN WATER
(ADAPTED FROM KOSTER ET AL. 2003). ......................................................................................10
TABLE 4 METHODS USED FOR THE DETECTION OF THE PATHOGENS OF CONCERN TO THE
MICRORISK PROJECT. ......................................................................................................................12
TABLE 5 STANDARDS FOR THE VALIDATION OF METHODS AND THE MONITORING OF
LABORATORY PERFORMANCE. ....................................................................................................14
TABLE 6 SOURCES OF BIOLOGICAL METHODS (FROM: ANON, 2004). ........................................16
TABLE 7 RESERVOIRS OF PATHOGENIC MICRO-ORGANISMS. ADAPTED FROM HURST ET
AL. 1997. ................................................................................................................................................18
TABLE 8 PREVALENCE OF ENTERIC PATHOGENS IN HUMANS, CATTLE, PIGS AND
POULTRY (OLSON, 2004A)...............................................................................................................19
4
TABLE 9 EXAMPLES OF PATHOGENS AND INDICATOR ORGANISMS COMMONLY FOUND IN
RAW SEWAGE. SOURCE: ADAPTED FROM YATES AND GERBA, 1998................................25
TABLE 10 CRYPTOSPORIDIUM AND GIARDIA OOCYSTS IN WASTE AND SURFACE WATERS.
AFTER ROSE (1990)............................................................................................................................26
TABLE 11 MANURES PRODUCED IN THE UK PER ANNUM, AND ESTIMATED
CAMPYLOBACTER CONTENT (STANFIELD AND GALE, 2002). .............................................27
TABLE 12. MANURES PRODUCED IN THE UK PER ANNUM, AND ESTIMATED E. COLI 0157
CONTENT (STANFIELD AND GALE, 2002). ..................................................................................27
TABLE 13 EFFECT OF TEMPERATURE ON INACTIVATION OF MICRO-ORGANISMS [DAYS] 29
TABLE 14 SUMMARY TABLE WITH CORRELATION TREND BETWEEN PARAMETER AND
PATHOGEN INACTIVATION RATE IN BRACKETS ....................................................................34
TABLE 15 SIZES OF SELECTED PATHOGENS ......................................................................................42
TABLE 16 ISOELECTRIC POINTS (PI) OF DIFFERENT PATHOGENS...............................................43
TABLE 17 SORPTION AND DESORPTION RATES FOR PATHOGENS IN SAND COLUMNS
[DAY-1]. ................................................................................................................................................44
TABLE 18 INFLUENCE OF MAJOR FACTORS ON THE SURVIVAL AND MIGRATION OF
MICRO-ORGANISMS IN THE SUBSURFACE. FROM PEDLEY ET AL. 2005. ...........................47
TABLE 19 PUBLICLY AVAILABLE VIRUS TRANSPORT CODES. FROM AZADPUR-KEELEY ET
AL. 2003. ................................................................................................................................................52
TABLE 20 OTHER VIRUS TRANSPORT CODES DEVELOPED FOR RESEARCH PURPOSES.
FROM AZADPUR-KEELEY ET AL. (2003).......................................................................................52
TABLE 21 MEAN CRYPTOSPORIDIUM AND GIARDIA DENSITIES IN THE RIVERS RHINE AND
MEUSE IN 1995 (FROM MEDEMA ET AL. 1996; VALUES CORRECTED FOR RECOVERY OF
DETECTION METHOD). ....................................................................................................................57
TABLE 22 SUMMARY OF CONCENTRATIONS OF SELECTED PATHOGENS IN WATER
BODIES. ................................................................................................................................................64
5
Background
Waterborne diseases
Waterborne disease remains one of the major health concerns in the World. Control of the
microbial quality of drinking-water should be a priority in all countries, given the immediate and
potentially devastating consequences of waterborne infectious diseases (WHO, 2004).
Diarrhoeal diseases, which are largely derived from poor water and sanitation account for 2.4
million deaths each year and contribute over 73 million Disability Adjusted Life Years (Prüss
and Havelaar, 2001). On a global scale, this places diarrhoeal disease as the sixth highest cause
of mortality and third in the list of morbidity. It is estimated that 5.7% of the global disease
burden is derived from poor water, sanitation and hygiene (Prüss et al. 2002). This health
burden is primarily borne by the populations in developing countries and by children. At present
estimates, one-sixth of humanity (1.1 billion people) lack access to any form of improved water
supply within 1 kilometre of their home and one-fifth of humanity (2.4 billion people) lack
access to some form of improved excreta disposal (WHO and UNICEF, 2000). In 2001,
infectious diseases accounted for an estimated 26% of deaths world-wide (Kindhauser, 2003).
In addition, social and environmental changes continue to result in new and re-emerging
waterborne pathogens. For example, climate change was estimated to be responsible for
approximately 2.4% of world-wide diarrhoea in 2000, 6% of malaria in some middle-income
countries and 7% of Dengue fever in some industrialised countries (Ashbolt, 2004).
This review is focussed on a selection of pathogens considered to be of high risk to human
health and which are considered to be of concern in source waters used for drinking water
supplies. These are: Campylobacter, E. coli 0157:H7, enteroviruses, norovirus,
Cryptosporidium, Giardia. Table 1 identifies the health significance of the pathogens of interest
in this report.
Table 1 Waterborne pathogens and their significance in water supplies. Source: WHO (2004).
Pathogen
Infectious dose
Campylobacter
jejuni, C. coli
Shown to vary but has been
caused by a few hundred
organisms (Percival et al.
2004). Most natural infections
probably require at least 104
organisms (Hunter 1997)
<100 organisms (Percival et
al. 2004). Consumption of
less than 50 organisms and
possibly as low as five
(Armstrong et al. 1996)
Difficult to assess but
generally thought that 1
infectious particle will infect
a susceptible host (Schiff et
al. 1984)
Median = 132 oocysts
(DuPont et al. 1995). ID50
was recalculated to be 87
oocysts (Fayer et al. 2000)
10-25 cysts (Rendtorff, 1954)
E.coli – enterohaemorrhagic
Enteroviruses
Cryptosporidiu
m parvum
Giardia
intestinalis
Norovirus
~10 viral particles (Bresee et
al. 2002)
Health
signifiCance
High
Persistence in
water
supplies
Moderate
Resistance
to chlorine
Low
Relative
infectivity
Moderate
Important
animal
source
Yes
High
Moderate
Low
High
Yes
High
Long
Moderate
High
No
High
Long
High
High
Yes
High
Moderate
High
High
Yes
High
Long
Moderate
High
Potentially
6
Water supply in Europe
Most water used for all purposes in Europe is abstracted from surface water sources, despite the
fact that the use of groundwater as a source of drinking water is often preferred because of its
generally good microbial quality in its natural state. However, evidence from around the World
has shown that groundwater may become rapidly contaminated if protective measures at the
point of abstraction are not well maintained. Further problems are caused by pollution in areas
where recharge of the source occurs, with persistent and mobile pollutants representing the
principal risks. Throughout the World, there is evidence of contaminated groundwater leading
to outbreaks of diseases and contributing to background endemic disease in situations where
groundwater sources used for drinking have become contaminated.
The importance of groundwater as a drinking water resource in Europe is highlighted in
Table 2 showing the proportion of groundwater in drinking water supplies in some
European countries. The data show that reliance upon groundwater varies considerably
between countries; for example, Norway takes only 13% of its drinking water from
groundwater sources, whereas Austria and Denmark are almost totally dependent upon
groundwater resources.
Table 2 Proportion of groundwater in drinking water supplies in selected European countries (EEA,
1999; UN ECE, 1999)
Country
Proportion
Country
Proportion
Austria
99%
Bulgaria
60%
Denmark
98%
Finland
57%
Hungary
95%
France
56%
Switzerland
83%
Greece
50%
Portugal
80%
Sweden
49%
Slovak
Republic
80%
Czech Republic 43%
Italy
80%
United
Kingdom
28%
Germany
72%
Spain
21%
Netherlands
68%
Norway
13%
Within countries the usage of groundwater may also vary substantially, depending on the terrain
and access to alternative water sources. For instance, in England and Wales, although the
national average for groundwater usage in 2003 was 33%, the Southern counties depend more
heavily
on
groundwater
than
the
Northern
counties
(http://www.dwi.gov.uk/pubs/annrep03/part1.htm).
Many waterborne disease outbreaks could be prevented by a good understanding and
management of drinking water sources for health. For example, pathogen contamination has
often been associated with simple deficiencies in sanitation but also with inadequate
understanding of the processes of attenuation of disease agents in the subsurface. This lack of
understanding easily leads to structures and practices overwhelming or by-passing attenuation.
Outbreaks of waterborne disease via public water supplies continue to be reported in developed
countries even though there is increased awareness of, and treatment for, pathogen
contamination (Herwaldt et al. 1992; Lisle and Rose, 1995; Moore et al. 1994; MacKenzie et al.
1994; Payment et al. 1997 and Howe et al. 2002).
7
The following sections of this review aim to provide a critical review of the analytical methods
relevant to the pathogens of interest; to identify the hosts, reservoirs and transmission pathways
of the pathogens; provide an insight into the contamination levels of source waters, and the
factors associated with those pathogens which may influence the contamination of water
sources.
8
Analytical Methods
For over 100 years our measure of the microbiological safety of water has relied upon the
isolation of a small group of non-pathogenic bacteria: the indicator bacteria. This group of
bacteria emerged as the foundation of water-related public health microbiology because many of
the pathogenic micro-organisms transmitted through water were either undiscovered, present in
very low numbers, difficult or impossible to culture, or too hazardous to be grown in a routine
water microbiology laboratory. The introduction of the indicator bacteria overcame many of
these obstacles and allowed the water utilities, environmental bodies and many other
organisations to rapidly, simply, and safely monitor the microbiological quality of water.
During the last 20 years, the reliability of the indicator bacteria as a means to assure the safety of
water has been increasingly challenged by water quality and public health microbiologists. In
support of this contention, there is a substantial library of publications that report the limited
correlation between the presence and concentration of indicator bacteria and the presence and
concentration of waterborne pathogens; in particular, demonstrating that indicator bacteria are
poor surrogates for protozoal and viral pathogens (Barrell et al. 2000; Berg and Metcalfe, 1978;
Griffin et al. 2001; Melnick and Gerba, 1982; Nwachuku et al. 2002; Payment et al. 1985;
Petrilli et al. 1974; Rose et al. 1986). Furthermore, several authors have shown that outbreaks of
waterborne disease have occurred despite indicator bacteria not being detected in the source
water (Barrell et al. 2000). These limitations have led several groups of workers to advocate the
routine testing of water for specific pathogens. Indeed, during the recent revision of the WHO
Guidelines for Drinking Water Quality, the WHO working committees created a list of reference
pathogens that would be used as part of a water quality monitoring and assessment programme.
Apart from the widely acknowledged limitation of indicator organisms as markers for the
presence or absence of waterborne pathogens, other factors are causing a shift towards water
quality monitoring using direct pathogen detection. These factors include:
•
•
•
Recent advances in the methods used for the isolation and detection of pathogens
in water.
The growth in our knowledge and understanding of waterborne pathogens.
The use of quantitative risk assessment models to calculate disease incidence in a
population exposed to a particular waterborne pathogen.
However, the shift towards the detection of pathogens once again draws attention to a
fundamental factor that limits the assessment of the microbial quality of waters, namely, the
often very low number of each micro-organism present. Thus, most of the analytical procedures
include three steps: concentration/enrichment, detection and quantification (Koster et al. 2003).
The routine detection of pathogens in water requires each one of these steps to be optimised,
since a significant development in one step may be offset by limitations in either one of the other
steps. Often, advances in diagnostic procedures are made in response to medical needs where
rapid identification and characterisation of a pathogen are the priority. The many advances in
biotechnology that have taken place to improve medical diagnostics are of particular benefit for
the detection of pathogens in water, but are of less value to the concentration/enrichment and
quantification steps. Consequently, for many of the methods described in this section,
concentration and quantification are limiting factors for the detection of pathogens in water.
The pathogens selected for analysis in this project belong to three very different groups: viruses,
bacteria and protozoa. Nevertheless, the procedures that have been described in the literature for
the detection of these pathogens in water are fundamentally very similar, with the differences
being confined to the specific diagnostic reagents used for the culture and detection of the
particular organism under investigation. Thus, a description of the methods used for the
detection of each pathogen will, inevitably, involve substantial repetition. A pragmatic approach
is to review the procedures used in water, irrespective of the organism under investigation, and
record the advantages and disadvantages of each. Koster et al. (2003) have published an
extremely detailed review of the analytical methods that are available, or being developed, for
microbiological water quality testing. The reader should refer to this review for a comprehensive
9
description of the methods. The pertinent details of each method relevant to this project - the
characteristics, limitations/disadvantages and applications - are contained in Table 3, which is
adapted from Koster et al. (2003).
Table 3 Methods for the detection of microbial contamination in water (adapted from Koster et al.
2003).
Method
Characteristics
advantages
Limitations
disadvantages
Application:
status quo and
future
perspective
of
• Cultivation media mostly
inexpensive.
• Easy to perform.
• Qualitative and quantitative
results obtainable.
• Differentiation
and
preliminary
identification
possible on selective solid
media.
• Detection
of
bacteria
occurring in low numbers
possible (in combination with
concentration techniques).
• Standardised (ISO,
CEN,
APHA)
methods
for
a
number of species
(groups).
• Improved
media
might be developed
in order to obtain
faster growth and to
increase sensitivity
and selectivity of the
assays.
Cultivation of
animal/human
viruses
• Several enteric viruses can
be propagated in cell culture
(a variety of cell lines have
been tested and used)
• Quantitation possible.
• Growth indicates infectivity.
• Time consuming.
• Not all bacteria of interest can
be cultivated.
• Large sample volumes cause
problems for some of the
methods.
• Does not detect viable but
non-culturable organisms.
• Selectivity of the detection of
certain indicators often not
sufficient
(false
positive
species).
• No information on infectivity
of a pathogen.
• Biosafety issues.
• Requires some level of
training
and
specialised
laboratories.
• Various cell lines may need to
be used for the detection of a
larger number of virus types.
• Biosafety issues.
• Several viruses cannot be
propagated on cell culure
Cultivation
protozoa
of
• Excystation in vitro can be
taken (to a certain extent) as
an indication of viability.
• Several protozoa can be
propagated in cell culture,
growth indicates infectivity
Immunological
detection
of
antigenic
structures
associated with
the
microorganisms
• Quantitative and qualitative
results regarding the number
of micro-organisms possible
(to a certain extent).
• Relatively specific for target
organism.
Immunomagnet
ic
separation
(IMS)
• Faster and more specific than
other concentration methods.
• Sound basis for other
detection methods (PCR, RTPCR, FACS, FISH) as well as
cultivation methods.
Cultivation
bacteria
• Time consuming.
• Sensitivity is low.
• Propagation
of
most
organisms in vitro using cell
cultures is poor.
• Not all protozoa of interest
can be cultivated.
• Biosafety issues.
• Often needs pre-cultivation
step which is time consuming.
• Lack of sensitivity.
• Selectivity can be a problem
due
to
cross-reacting
antibodies.
• Without
pre-cultivation,
currently no discrimination
between viable and non-viable
micro-organisms.
• No information on infectivity
of a pathogen.
• Sensitivity,
robustness,
consistency can be affected by
environmental conditions.
• Selectivity can be a problem
due
to
cross-reacting
antibodies.
• No information on infectivity
of a pathogen.
• Standardised (ISO,
CEN,
APHA)
methods
for
a
number of species
(groups).
• New cell lines are
being developed and
new
media
formulation
may
increase sensitivity.
• Application
is
limited due to low
sensitivity. New cell
lines and media may
improve sensitivity.
• Assays
allow
standardisation and
automation.
10
Method
Characteristics
advantages
Limitations
disadvantages
Application:
status quo and
future
perspective
Polymerase
chain reaction
(PCR)
• In principle, highly sensitive
(but see limitations).
• Selective.
• Specific.
• Can detect non-culturable
organisms.
• Faster
than
cultivation
methods (3-4 hours).
• Sound basis for further
analysis of nucleic acids
(sequencing, RFLP, RAPD).
• Currently
no
standardisation.
• Potential
for
automation.
• Potential
for
quantitation in real
time PCR.
RT-PCR
• As PCR.
• Good indication of living
organisms with mRNA as
target.
• Can provide information on
pathogenic potential of an
organism when mRNA of a
virulence gene is assayed.
Flow
cytometry,
fluorescenceactivated cell
sorting (FACS)
• Faster
than
cultivation
methods.
• Detection of non-culturable
organisms.
Fluorescence in
situ
hybridisation
(FISH)
• Faster
than
cultivation
methods.
• No pre-cultivation needed.
• Detection of non-culturable
organisms.
• Can detect individual cells
when rRNA is target.
• Different
(multicolour)
fluorescent
labels
allow
detection of different microorganisms.
• Can be used in combination
with machines that do
automated scanning of filter
surfaces
for
fluorescent
objects.
• Faster
than
cultivation
methods.
• Excellent
tool
for
differentiation of strains or
isolates within a species.
• Limited reliability (at present
the detection of an individual
microbe cannot be guaranteed
due to inconsistencies in the
performance of the technique).
• Sufficient quantity of nucleic
acids from the targeted microorganism has to be recovered.
• Negative affect by certain
environmental conditions.
• Basic procedure does not
allow quantitation of the
number
of
amplifiable
DNA/RNA fragments.
• At present no discrimination
between viable and non-viable
micro-organisms.
• No information on infectivity
of a pathogen.
• As
PCR
(except
discrimination between viable
and
non-viable
microorganisms with mRNA as
target).
• Extraction of detectable levels
of intact RNA molecules is
problematic due to their
instability.
• No information on infectivity
of pathogen.
• Expensive technology.
• Limited reliability for the
detection of micro-organisms
that are present in extremely
low concentrations.
• Lack of sensitivity with
chromosomal genes or mRNA
as target.
• Detection
is
strictly
taxonomic.
• Differentiation between living
and dead cells is often difficult.
• Not applicable to detect 1
indicator per 100ml without
concentration/filtration.
Molecular
fingerprinting
(ribotyping,
RFLP, RAPD,
AP-PCR)
• Currently
no
standardisation.
• Potential
for
automation.
• Potential
for
quantitation.
• Potential
automation.
for
• At present no discrimination
between viable and non-viable
micro-organisms.
• RAPD requires the use of pure
isolates.
11
Method
Characteristics
advantages
Limitations
disadvantages
Application:
status quo and
future
perspective
DNA chip array
• Micro-manufacturing
techniques allow testing of up
to several thousand sequences
in one assay on a single chip.
• Sensitive, selective and
specific to the desired level to
detect groups of organisms or
(sub)-species, respectively.
• Fast (2-4 hours)
• Immunoaffinity step to bind
micro-organisms to surfaces;
detection by laser excitation
of the bound fluorescent
antibodies,
acoustogravimetric
wave
transduction,
or
surface
plasmon resonance.
• Rapid, but depends on
culturable micro-organisms.
• At present very cost intensive.
• Highly trained personnel
needed.
• Absolute quantitation may be
problematic.
• Substantial concentration of
samples required
• Technique not yet
widely available.
Biosensors
• Currently unable to
discriminate
between viable and
non-viable
microorganisms.
Not every method has been used for detecting each pathogen; indeed, some methods would be
particularly unsuitable for the detection of all pathogens. Table 4 is a matrix of the methods of
detection and the pathogens of interest to this project showing which methods have been used
for each pathogen. The literature is heavily populated with references describing subtle
variations to each of the principal methods, which may improve the sensitivity or specificity of
detection in particular environmental matrix. Table 4 contains only selected references
demonstrating the use of each method.
Table 4 Methods used for the detection of the pathogens of concern to the MICRORISK project.
Crypto - Giardia
sporidium
Cultivation
bacteria
of
Cultivation
viruses
of
Cultivation
protozoa
of
Immunological
detection
of
antigenic structures
Campylo
– bacter
E.coli
O157
Anon
(2002b);
Anon (2004);
Hanninen et
al. (2003);
Percival et al.
(2004)
Anon (2002a);
Anon (2004);
March
and
Ratnam
(1986)
Entero - Noro viruses
viruses
Percival et
al. (2004);
CEN (2000)
Carey et al.
(2004);
QuinteroBetancourt et
al.
(2002;
2003).
Carey et al.
(2004);
Nieminski et
al.
(1995);
QuinteroBetancourt et
al.
(2002;
2003).
Nieminski et
al.
(1995);
QuinteroBetancourt et
al. (2003).
Koster et al.
(2003)
12
Crypto - Giardia
sporidium
Immunomagnetic
separation
Polymerase
reaction
chain
Carey et al.
(2004);
MassanetNicolau
(2003);
QuinteroBetancourt et
al.
(2002;
2003).
Carey et al.
(2004); Guy et
al.
(2003);
Nichols et al.
(2003)
Campylo
– bacter
MassanetNicolau
(2003);
QuinteroBetancourt et
al. (2003).
Caccio, S.M.
et al. (2003);
Guy et al.
(2003)
E.coli
O157
Entero - Noro viruses
viruses
Tomoyasu
(1998)
Moreno et al.
(2003); Oyofo
and
Rollins
(1993);
Waage et al.
(1999);
Fukushima et
al. (2003)
Reverse
transcription PCR
Ibekwe et al.
(2002);
Fukushima et
al. (2003)
Beuret
(2003);
Borchardt et
al. (2003);
Fout et al.
(2003)
Flow
cytometry,
fluorescence
activated
cell
sorting
Fluorescence in-situ
hybridisation
Molecular
fingerprinting
Medema et al.
(1998)
Medema et al.
(1998)
Graczyk et al.
(2003)
Nichols et al.
(2003)
Graczyk et al.
(2003)
DNA chip array
Straub et al.
(2002)
Beuret
(2003);
Lamothe et
al. (2003);
Parshionikar
et
al.
(2003).
Moreno et al.
(2003)
Hanninen et
al. (2003)
Methodological considerations
Assessment of the risk of infection from waterborne pathogens requires accurate determinations
of microbial occurrence, concentration, viability and infectivity, and human dose-response data
(LeChevallier et al. 2003). Each of the methods listed above has limitations in one or more of
the criteria; for example, nucleic acid and antibody-based methods do not readily provide
information about the concentration, viability and infectivity of the pathogen, whereas culture
methods can be used only for the relatively small group of pathogens that are capable of growth
in culture. Furthermore, the recovery rates of many culture methods may be very low, leading to
a significant underestimate of pathogen numbers. It is important when selecting the method of
analysis to balance the advantages and disadvantages of each in terms of the outputs that are
required.
An important consideration for any multi-centre project is that the methods of analysis must be
available to all the laboratories, and be sufficiently detailed in their scope to ensure comparable
results from the participating laboratories. Therefore, wherever possible, international standard
methods should be used, supported by regular monitoring of inter-laboratory performance. In
any case, participating laboratories should provide their Quality Assurance/Quality Control data
on performance charaterisitics of the method they are using.
Standard methods of analysis are published by several organisations (for example, ISO, CEN,
APHA) and there are many supporting standards for the validation of methods and the
monitoring of laboratory performance (Table 5).
13
Table 5 Standards for the validation of methods and the monitoring of laboratory performance.
Topic
Laboratory AQC
and method
validation
Cryptosporidium
and Giardia
Title
Reference/
Publisher
General requirements
for the competence of
testing and calibration
laboratories.
BS EN ISO/IEC
17025:2000
Accuracy (trueness and
precision) of
measurement methods
and results.
BS ISO 5725
Proficiency testing by
interlaboratory
comparison.
Development and
operation of
proficiency testing
schemes
Quality assurance /
Quality control
PD6644-1:1999
ISO/IEC Guide 431:1997
Quality Assurance:
Principles and practice
in the microbiology
laboratory
Public Health Laboratory
Service. ISBN
0901144452 (Snell,
Brown and Roberts, eds)
Microbiological
Analysis of Food and
Water: Guidelines for
Quality Assurance
The Microbiology of
Drinking Water (2002)
- Part 3 - Practices and
procedures for
laboratories
Elsevier. ISBN
0444502033 (Lightfoot
and Maier, eds)
Water Quality:
Isolation and
enumeration of
Cryptosporidium
oocycts and Giardia
cysts from water
APHA, AWWA, WEF.
Standard Methods for the
Examination of Water
and Wastewater. Section
9020
UK Environment
Agency. The documents
can be downloaded from
their website
(www.environmentagency.gov.uk)
ISO/DIS 31598
Comments
A comprehensive quality
standard covering all
aspects of laboratory
management and control.
The standard makes
reference to many other
standards that may be of
interest to laboratories.
The standard is published
in six parts, each one
dealing with a different
aspect of testing the
accuracy of measurement
methods and results
Standard Methods section
9020 contains three
sections under QA/QC:
Introduction;
Intralaboratory Quality
Control Guidelines;
Interlaboratory Quality
Control.
Although written for
clinical laboratories, the
book includes sections that
are relevant to all
microbiology laboratories.
The book covers all aspects
of quality assurance for
microbiological analysis of
water.
This document provides a
full review of the practices
and procedures that should
be operating in a
microbiology laboratory
carrying out water quality
analysis.
The document is available
as a draft for public
comment.
14
Topic
Title
Reference/
Publisher
Comments
Part B of section 9711
APHA, AWWA, WEF.
Standard Methods for the describes the methods for
the detection and
Examination of Water
and Wastewater. Section enumeration of Giardia and
Cryptosporidium.
9711
The International Organization for Standardization (ISO) has not published
methods for the isolation of E.coli O157 from water; however, it has published
methods for its isolation from food and animal feeding stuffs (BS EN ISO 16654:
2001).
Part F of section 9260
Detection of
APHA, AWWA, WEF.
Pathogenic Bacteria
Standard Methods for the describes method for the
isolation and identification
Examination of Water
and Wastewater. Section of pathogenic E.coli
9260
Section F describes
UK Environment
The Microbiology of
Drinking Water (2002) Agency. The documents methods for the selective
can be downloaded from enrichment, isolation and
- Part 4 - Methods for
identification of E.coli.
their website
the isolation and
The methods rely on
(www.environmentenumeration of
growth of the bacterium in
agency.gov.uk)
coliform bacteria and
culture and its
Escherichia coli
identification by
(including E.coli
biochemical and
O157:H7)
immunological methods.
The International Organization for Standardization (ISO) has not published
methods for the isolation of Campylobacter from water; however, it has published
methods for the isolation of thermotolerant Campylobacter from food and animal
feeding stuffs (BS 5763-17: 1996; ISO 10272: 1995).
Part G of section 9260
Detection of
APHA, AWWA, WEF.
Pathogenic Bacteria
Standard Methods for the describes method for the
isolation and identification
Examination of Water
and Wastewater. Section of Campylobacter jejuni
9260
Section C describes
UK Environment
The Microbiology of
Drinking Water (2002) Agency. The documents methods for the selective
- Part 10 - Methods for can be downloaded from isolation of thermophilic
Campylobacter species by
their website
the isolation of
selective enrichment. The
(www.environmentYersinia, Vibrio and
methods rely on growth of
agency.gov.uk)
Campylobacter by
the bacteria in culture and
selective enrichment
their identification by
biochemical,
morphological and
physiological tests.
Section 9510 describes
Detection of enteric
APHA, AWWA, WEF.
viruses
Standard Methods for the several methods for the
recovery of enteroviruses
Examination of Water
and Wastewater. Section from water. Part G
describes the assay and
9510
identification methods.
Water Quality.
BS EN14486
The method has been
Detection of human
published as a draft for
enteroviruses by
public comment. The assay
monolayer plaque
methods described are
assay.
based on the growth of the
viruses in cell culture.
Pathogenic Protozoa
Escherichia coli
O157
Campylobacter
Enteroviruses
15
Noroviruses
There are no international standard methods for the detection of noroviruses in
water.
Often, international standard methods of analysis are not available for waterborne pathogens and
it is necessary to use alternative sources of standardised methods. Table 6 provides a list of
sources of methods.
Table 6 Sources of biological methods (from: Anon, 2004).
Name
National Environmental
Methods Index (NEMI)
U.S. EPA microbiology
methods
USDA/FSIS
Microbiology laboratory
guidebook
ICR
Microbial
laboratory manual
Publisher
EPA, USGS
Reference
www.nemi.gov
EPA
www.epa.gov/microbes/
USDA Food safety
and
inspection
service
EPA
Office
of
research
and
development
Occupational safety and OSHA
health
administration
methods
National institutes for NIOSH
occupational safety and
health methods
Public
Standard methods for the American
examination of water Health Association
and American Water
and wastewater
Works Association
Annual book of ASTM ASTM International
standards
Applied
and American society for
microbiology
Environmental
Microbiology
Journal
of
Clinical American society for
Microbiology
microbiology
Standardised analytical EPA
methods for use during
homeland security events
Environment
The microbiology of UK
drinking water. Methods Agency
for the examination of
waters and associated
materials
ISO
International
Organization
for
Standardization
www.fsis.usda.gov/ophs/microlab/
mlgbook.htm
www.epa.gov/nerlcwww/icrmicro.
pdf
www.oshaslc.gov/dts/sltc/methods/toc.html
www.cdc.gov/niosh/nmam/
www.apha.org
www.awwa.org
ISBN: 0875532357
www.astm.org
www.asm.org
www.asm.org
www.epa.gov/ordnhsrc/pubs/repor
tSAM092904.pdf
www.environment-agency.gov.uk
www.iso.org
The methods of analysis using the amplification and/or detection of nucleic acids and the
immunological detection of antigenic structures have been used for many years and are routine
in many clinical diagnostic laboratories. Diagnostic kits using these methods are available
16
commercially for a broad-range of pathogens, and several have been tested for their application
to environmental samples. The number of publications describing the use of nucleic acid or
immunological methods for the detection of pathogens in water is already very large and
continues to grow with the appearance of each new journal. Nevertheless, the validation
procedures in these papers are often limited and sometimes non-existent. This is a significant
weakness of the methods that must be addressed as part of any multi-centre study, particularly
for the detection of those pathogens for which no other methods of analysis are available.
Conclusions
During the last 20 years, advances in analytical methods for the detection of micro-organisms
have opened the possibility of rapidly and simply ascertaining the presence of pathogens in
water. The application of these methods to the analysis of waterborne pathogens has produced
an overwhelming number of publications; however, the underlying processes are fundamentally
very similar and change only between the target of the analytical method (whole organism,
antigenic structure or nucleic acid).
Assessment of the risk of infection from waterborne pathogens requires accurate determinations
of microbial occurrence, concentration, viability and infectivity, and human dose-response data.
As shown above, there is no single method that meets all these requirements, and most methods
have weaknesses in at least two. Thus, method selection will require a careful assessment of the
features of each method with respect to the quality and quantity of data that is required, the
nature of the organism, the nature of the environment that is being sampled, the cost, and the
requirement for reproducibility between laboratories. Ultimately, the chosen method will
represent a compromise, and it is important to understand the limitations of the method as they
impact upon the risk assessment.
Methods of analysis should be fully validated before they are used. This applies to any method,
including international standard methods, but it is particularly important for the methods
developed recently as a result of advances in biotechnology. Although methods based upon
nucleic acid detection and characterisation have been widely published, it is very uncommon for
the method to have been fully validated. Ease of validation is also an important consideration
for multi-centre studies where comparability of data sets is critical. Consequently, it may be
prudent, wherever possible, to use international standard methods of analysis and use more
innovative techniques only where standard methods are not available.
The concentration of pathogens in water is often very low. Concentration and/or enrichment of
the pathogen from large volumes of water, sometimes thousands of litres, is essential before it
can be detected using any of the analytical methods. Consequently, it is important to the success
of the monitoring programme that standard methods of sampling and sample processing are used
by the participating laboratories.
17
Review of basic knowledge of the sources and
occurrence of chosen pathogens.
The relative significance of the different sources of occurrence of pathogens at a specific water
site is determined by a combination of factors: (1) the contamination level of these sources, (2)
the magnitude of these sources, (3) the persistence of the pathogen, (4) their transport behaviour
from the source to the specific site and finally, (5) their resistance against treatment processes.
Knowledge of these characteristics and about the health outcome after infection allows the
appraisal of the health significance of the pathogen. The source of the pathogen, including the
potential reservoir, and the pathogen loadings in the source and the water body are of importance
in assessing the risk posed by the water body that these pathogens are found in. The pathogens
of particular interest in this project have been selected because they are considered of high health
significance, and are an issue in source waters rather than those such as Legionella which are,
for example, a particular problem of re-growth in the distribution system.
Reservoirs of chosen pathogens
The reservoirs for the pathogenic micro organisms found in environmental waters can be either
humans, animals or the environment itself (Table 7).
Table 7 Reservoirs of pathogenic micro-organisms. Adapted from Hurst et al. 1997.
Reservoir
Human
Human and animal
Animal
Environmental
Disease
Cholera
Encephalitis
Entamoeba
Gastroenteritis
Hepatitis
Meningitis
Campylobacteriosis
Cryptosporidiosis
Giardiasis
Severe gastroenteritis
Leptospirosis
Encephalitis
Cholera
Legionellosis
Causative genus
Vibrio
Enterovirus
Entamoeba
Astrovirus,
Norovirus,
Coronavirus, Rotavirus
Calicivirus, Hepatovirus
Enterovirus
Campylobacter
Cryptosporidium
Giardia
E. coli 0157:H7
Leptospira
Naegleria
Vibrio
Legionella
Most microbial waterborne pathogens of concern originate in the enteric tracts of humans or
animals and enter the aquatic environment via faecal contamination either directly through runoff or from sewage or manure.
Potential Sources of contamination
Overview
Source waters are vulnerable to contamination from many sources. Potential contaminants
include products from agriculture and animal husbandry, chemicals and micro-organisms in runoff from agricultural land, chemicals in industrial discharges, and nutrients and pathogens from
domestic sewage. Source waters, particularly surface waters, are often used for many purposes
other than water supply, such as irrigation, recreation, waste discharge and transport, and these
may affect the water quality. Point source discharges, such as sewage management systems and
18
agricultural facilities, including manure, manure-processing and slaughterhouse wastewater, are
normally diluted and carried away from the source. However, multiple discharges along the
course of a river can result in increasing levels of contamination downstream and it has been
shown that many rivers in Europe are significantly contaminated with microbes (EEA, 2003),
arising from municipal wastewater and/or animal husbandry, which are of public health concern.
Sewage treatment plants are an obvious high risk source of pathogens, both in terms of numbers
and strains of pathogens likely to be infectious to humans. Sewage treatment plants are a
concern because they can release significant amounts of poorly treated effluent during periods of
high rainfall or plant failure and they can widely distribute pathogens in the environment
through sewage sludge use as fertilisers.
Leaking sewage (pipe) systems are currently being studied in detail by an ongoing EU cluster
called Citynet (citynet.unife.it). This series of projects tries to cope with the already occurring
incidents and the potential problems linked with leaking sewage pipes in urban areas because the
amounts lost from the pipe system are usually unknown and largely depend on the quality and
maintenance of the system.
Sewer overflows are installed as part of the sewerage system to prevent wastewater from
backing up in domestic properties. The impact of these in receiving waters largely depends on
the total amount of contaminants discharged, their location in the river system and the frequency
of occurrence. During extended or heavy rain, flow can enter the system by illegal storm water
connections or seep through cracks in pipes.
A further concern is on-site sewage management facilities which serve single residences in
unsewered areas. The basic function of these systems is to treat all the wastewater produced by
a household and distribute it to adjacent land. There is a broad range of on-site systems available
including septic tanks with associated absorption fields and composting toilets.
As well as sewerage systems, other sources of faecal contamination which can pose a threat to a
water source include: storm water, or urban runoff; accumulation of pathogens in sediment;
swimming pool water and water treatment plant discharges and feral animals. Advances in
source tracking techniques (for review of techniques see Meays et al. 2004; Pond et al. 2004)
which differentiate between animal and human sources of faecal pollution will allow more
precise information on the sources of contamination and will assist water resource managers to
develop strategies to protect source waters and thus reduce public health risks from these waters.
The following sections provide more detailed information on the sources and health implications
of the pathogens selected for this study.
Hosts of pathogens and health implications
As infected livestock have considerable potential for contaminating aquatic environments,
agricultural practices are an important source of contamination (Table 8) (Carey et al. 2004;
Lack, 1999).
Table 8 Prevalence of enteric pathogens in humans, cattle, pigs and poultry (Olson, 2004a)
E.coli O157
Campylobacter
Giardia lamblia
Cryptosporidium
hominis/parvum*
Human
[%]
1
1
1-5
1
Cattle
[%]
16
1
10-100
1-100
Pigs
[%]
0.4
2
1-20
0-10
Poultry
[%]
1.3
100
0
0*
* C. meleagridis is found in turkeys
Particularly during heavy rainfall events, it is likely that run-off of animal manures and soil
loads occur from grazing lands. Cattle faecal matter has been shown to be a significant source of
Cryptosporidium oocysts for instance. Calves and lambs are known to produce prolific numbers
of oocysts, with as many as 10 million oocysts per gram of faeces being excreted from infected
19
calves (Fayer et al. 1989). Waterborne outbreaks of disease have been reported both from
human effluent and cattle throughout the World.
Agricultural activities can contribute to water pollution, not only by the production of animal
contamination but also by disturbance of vegetation near waterways. Such disturbance removes
the natural barriers to pollutants entering the water.
Biosolids or wastewater solids produced from the treatment of sewage must be disposed of.
There are a number of options for disposal, one of which is disposal to receiving waters.
Research has shown that Giardia cysts are still capable of detection in biosolids for up to six
months at high levels (McInnes et al. 1997).
Cryptosporidium
Cryptosporidium is a significant cause of waterborne outbreaks of diarrhoeal diseases. The
Centre for Disease Control and Prevention in Atlanta, USA attributed 71% of waterborne
disease outbreaks in 1993 and 1994 to Cryptosporidium parvum (C. parvum) and Giardia
lamblia, which cause cryptosporidiosis and giardiasis, respectively (Gostin et al. 2000). Attack
rates of cryptosporidiosis in these outbreaks are about 40% for the population at risk, as
compared to 5-10% for giardiasis (Smith and Rose, 1990). Approximately 50 drinking-water
related outbreaks due to Cryptosporidium were reported between 1984 and 1999, mostly in
North America, the UK, and Japan (Fayer et al. 2000) where detection and monitoring systems
were in place. The larger outbreaks have been recorded in Texas (2006 persons affected),
Georgia (12,960 persons), Oregon (15,000 persons), Ontario (>1000 persons), British Columbia
(14,500 persons), British Columbia (2097 persons), Japan (>9000 persons), and the UK (14,500
persons) (Fayer, 2004). Cryptosporidium was responsible for the largest water-borne disease
outbreak ever recorded for any pathogen, resulting in cryptosporidiosis in approximately
403,000 persons in Milwaukee, Wisconsin, USA in the spring of 1993, due to Cryptosporidium
with no species identified (MacKenzie et al. 1994). Subsequent studies indicate that this
outbreak was caused by C. hominis (Zhou et al. 2003). Based on death certificate records for two
years following the outbreak, cryptosporidiosis-associated deaths were reported for 54 residents
(Hoxie et al. 1997).
Cryptosporidium infections occur predominantly in very young (neonate) animals, only humans
seem to be susceptible at any time in their lives. It is a particular problem in those people with a
reduced immune system, especially those with Acquired Immune Deficiency Syndrome
associated with HIV infections (Hunter and Nichols, 2002).
Cryptosporidium infects both farmed and wild animal hosts including fish, snakes, birds, mice,
rats, cats, dogs, squirrels, deer, horses, pigs, sheep, cattle and others (Tzipori, 1983; Fayer and
Ungar, 1986). In New Zealand, the highest sample prevalence was in areas of intensive livestock
farming (Ionas et al. 1998). Hoogenboezem et al. (2001) report 90% of newborn veal calves in
the Netherlands to be positive for Cryptosporidium and Giardia. Some species (e.g., rats, mice,
guinea pigs) appear to have innate resistance as their infections are asymptomatic, whereas
others (e.g., ruminants) are, as are humans, susceptible to disease (Tzipori, 1983).
Molecular typing tools have shown that two genotypes of Cryptosporidium parvum are
responsible for outbreaks of waterborne diarrhoeal disease (Peng et al. 1997; Sulaiman et al.
1998). The human genotype (genotype 1; C. hominis) parasites have so far been found only in
humans, whereas the bovine genotype (genotype 2; C. parvum) parasites have been found in
farm animals and humans (Fayer et al. 2000). Detection of genotype 1 is therefore indicative of
human contamination of the water body, whereas detection of genotype 2 could be either from
an animal or human source. Drinking-water borne outbreaks have been associated with both
genotypes, and descriptive data have shown the possibility of both human and animal sources of
contamination in source waters (Casemore, 1998; Dolej et al. 2000). Recent research has shown
that the population structure of C. parvum and C. hominis is apparently more complicated than
previously suggested, with the likely existence of both clonal and panmictic populations. Thus,
the transmission of C. parvum (genotype 2) in humans is shown to vary in different areas, with
zoonotic transmission important in certain places and anthroponotic transmission in others.
Apart from livestock and the cattle genotypes of C. parvum it has recently been suggested that
20
the role of other mammals and birds in zoonotic transmission of Cryptosporidium is uncertain.
It is known that humans can be infected by other species of Cryptosporidium, such as the
previously presumed avian-specific species C. meleagridis, but the prevalence of the various
species and genotypes of Cryptosporidium is unknown and the frequency of cross-contamination
is also unknown (Monis and Thompson, 2003).
The use of molecular tools has also led to the identification of geographic and temporal
differences in the transmission of C. parvum and C. hominis, and better appreciation of the
public health importance of other Cryptosporidium species/genotypes and the frequency of
infections with mixed genotypes or subtypes (Xiao and Ryan, 2004).
Cryptosporidium infection is transmitted through animal-to-person contact or person-to-person
contact, or through contact with fecally-contaminated surfaces, as well as via ingestion of
fecally-contaminated food or water. Regan et al. (1996) for example, implicated an outside
garden hose that had probably lain in fecally-contaminated grass, and was subsequently used to
fill drinking water coolers at a day camp in a case of cryptosporidiosis. Outbreaks associated
with fecally-contaminated recreational waters (Bell et al. 1993; McAnulty et al. 1994), day care
centres (Alpert et al. 1984), infected farm animals (Miron et al. 1991; Lengerich et al. 1993)
have also been recorded. Laboratory research animals have been implicated as sources of
infection (e.g., Anderson et al. 1982), and some "traveller's diarrhoea" is also likely attributable
to Cryptosporidium (Ma et al. 1985; Soave and Ma, 1985). No transmission from household pets
to humans has been proven, but there are suspicions of such episodes (Juranek, 1995). Although
contaminated food is considered a source of Cryptosporidium, there seem to be few documented
incidents. There was one outbreak among individuals who drank fresh-pressed apple cider at a
county fair. The cider was pressed from orchard-collected apples, including some fruit from the
ground, apparently contaminated with animal faeces (Millard et al. 1994). Inadvertent faecal
contamination of foodstuffs is implicated in many instances of food borne illness. It is
reasonable to surmise that infected food handlers could also unwittingly transmit
Cryptosporidium infection by contaminating beverages, salad greens or other uncooked foods
with oocysts. Cooked foods would be safe unless re-contaminated, because the oocysts are heatsensitive. Juranek (1995) observes that ~50 % of dairy calves shed oocysts, and the parasite is
present >90% of dairy farms. This implies that ingestion of unpasteurised milk could lead to
cryptosporidiosis.
Sewage/wastewater treatment has been shown to decrease oocyst content, but oocysts remain in
the treated effluent, suggesting that sewage discharge may be a significant source of oocysts in
the environment (Rose, 1990). The identification of isolates from open waters has revealed a
great diversity of unknown genotypes, of which many were later isolated from animal sources
(Perz and Le Blancq, 2001; Xiao et al. 2001; Xiao et al. 2002).
On the basis of data collected on untreated sewage water and manure, it can be calculated that
annually 1.87 x 1016 Cryptosporidium oocysts are produced in the Netherlands; 84% are from
liquid calf manure, 13% from households and 2.3% from commercial egg layers manure
(Hoogenboezem et al. 2001). Although calf manure applied to land is considered to be a
potentially significant source of Cryptosporidium and Giardia, it is not known, however, what
percentage of the oocysts from animal manure reach the surface water as a result of run-off of
manure applied to the land. It is thought to be only a small fraction. In addition, Hoogenboezem
et al. (2001) report that treated wastewater from cattle, pig and poultry slaughterhouses do not
make a significant contribution to the discharge of Cryptosporidium and Giardia in surface
waters. However, this may not be so with other animals or in other areas.
Although agricultural sources (e.g. runoff from dairies, grazing lands) are clearly a major
concern (Table 11 and Table 12) it has recently been suggested that the source of infections with
Giardia and Cryptosporidium may be more often from other humans, rather than from cattle via
pasture run-off, than first thought (Olson et al. 2004b). Certainly, in the USA and in Canada,
cattle have not been conclusively identified as the source of any waterborne outbreak of
cryptosporidiosis (Olson et al. 2004b). However, this needs further investigation in other
geographical locations since there are a many reports of cattle, sheep and other livestock and
wildlife infected with Cryptosporidium and associated water bodies also being contaminated
(see for example, Sturdee et al. 2003). One exception is the waterborne outbreak in Cranbrook,
21
British Columbia, Canada, where oocysts of the bovine genotype have been identified (Fayer et
al. 2000).
Giardia
Giardia has been reported as the most common cause of protozoan diarrhoeal illness worldwide
(Farthing, 1989; Adam, 1991). Between 1971 and 1994, more than 25,000 cases of giardiasis
were recorded in the USA (Craun, 1986; Anon, 1993; 1996).
Infections have been documented from drinking contaminated water from streams, rivers,
springs and ponds, infected household and day care contacts, especially children in
diapers/nappies; swimming in untreated surface water, such as wading pools, ponds, rivers,
streams or lakes private water systems (wells or springs) that are not correctly installed or
maintained. Many studies have shown the presence of Giardia in surface water and groundwater
(LeChevallier and Norton, 1996; Hancock et al. 1998).
There is a large body of evidence demonstrating the occurrence of G. lamblia in human
wastewater as well as in animal wastes (Sykora et al. 1988; 1991). Some animals are believed to
serve as reservoirs for human pathogenic strains, with much attention being given to beavers and
other aquatic animals. Giardia duodenalis, the species that infects humans, consists of seven
genotypes. Among them, genotype AI is found in humans, livestock, dogs, cats, beavers and
other animals; genotype AII is found only in humans; and genotype B is found in humans,
beavers, dogs, muskrats and other animals. Transmission from humans to beavers, dogs and
muskrats suggests some Giardia are zoonotic, and similar gene sequences among isolates
support this possibility. Genotypes C and D are found primarily in canids, and genotypes E, F
and G, found primarily in hoofed livestock, cats and rats, respectively, have not been found in
human infections (Fayer, 2004). Evidence to support the zoonotic transmission of Giardia is very
strong, but how frequent such transmission occurs and under what circumstances, has yet to be
determined. It is clear that aquatic animals, domestic dogs and cats, and cattle may serve as
sources of measurable cysts in surface waters. Water sources near farms are particularly
vulnerable to Giardia contamination. However, only a single genotype has been unequivocally
demonstrated to infect both humans and animals (Monis and Thompson, 2003) and according to
Thompson (2004) zoonotic origin for waterborne outbreaks of Giardia infection appears to be
uncommon. Similarly, livestock are unlikely to be an important source of infection in humans.
The greatest risk of zoonotic transmission appears to be from companion animals such as dogs
and cats, although further studies are required in different endemic foci in order to determine the
frequency of such transmission (Thompson, 2004). Most outbreaks of giardiasis have been
linked to consumption of water contaminated by human sewage (Thompson et al. 2000). There
are a number of reports showing contamination of surface water by discharges of untreated and
treated domestic sewage (Sykora et al. 1991; Rose et al. 1986).
Campylobacter
Campylobacter is considered the most important bacterial agent in waterborne diseases in many
European countries (Stenström et al. 1994; Furtado et al. 1998) - a large number of outbreaks of
Campylobacter have been reported in Sweden for example, involving over 6000 individuals
(Furtado et al. 1998).
Most species of Campylobacter are adapted to the intestinal tract of warm-blooded animals. It is
now known that Campylobacters are widespread in the environment, and that some strains of
Campylobacter isolated from patients stools can be found in livestock, poultry, wild birds, farms,
sewage and surface waters (Jones, 2001; Levesque et al. 2000; Park, 2002). Campylobacter does
not thrive in foodstuffs or or water, owing to its very special habitat requirements (+42oC, microaerophilic).
Due to difficulties with accurate species identification of Campylobacter, clinical laboratories
usually make no distinction between C. jejuni and C. coli. Relatively few studies have been
conducted aiming at species identification of patients' isolates. However, the general idea is that
C. jejuni predominates, accounting for 80–90% of all cases, and that 5–10% are due to C. coli,
when the diagnosis is based on culture-selective media (Nachamkin et al. 2000). Other human
22
pathogen types such as C. laridis, C. fetus and C. uppsaliensis are normally not detected at the
clinical laboratories, depending on the methods used. Apart from the consumption of
contaminated food and water, there are no clearly defined routes for the transfer of
Campylobacters from the environment to the consumer (Jones, 2001). Farmed animals may be a
source of these organisms and dairy cows have been reported as playing a significant role as a
reservoir of Campylobacter subtypes that can cause human disease (Stanley and Jones, 2003).
Several outbreaks in the Nordic countries have also suspected seagulls and other waterfowls to
play a role in the contamination of surface water or an open reservoir.
It has been indicated that in general, issues with Campylobacter in drinking waters tend to be a
post-treatment problem, due to situations such as broken sewers. Outbreaks of
campylobacteriosis have not been associated with properly-operated disinfected public water
supplies (G. Stanfield, Pers. Comm.). Waterborne Campylobacter infections tend to be
associated with private supplies, or with public supplies lacking disinfection or adequate
treatment, though the association is often unproven (G. Stanfield, Pers. Comm). However,
Campylobacters have been isolated from rivers (Arvantidou et al. 1996; Obiri-Danso and Jones,
1999), lakes (Arvantidou et al. 1996), groundwater (Savill et al. 2001) as well as drinking water
(Alary and Nadeau, 1990; Savill et al. 2001; Vogt et al. 1982). The occurrence of the organisms
in surface waters has also proved to be strongly dependent on rainfall, water temperature and the
presence of waterfowl (WHO, 2004).
No indicator bacteria have been found during investigations of some of the larger Swedish
waterborne outbreaks. This has sometimes made the investigation more difficult, particularly
where there was no good explanation of the source of the outbreak, even if it was proven
epidemiologically to be waterborne. There were no suspicions of sewage contamination of the
water (neither source or tapwater) in any of these outbreaks. In three of the outbreaks there
might have been contamination from birds through the sand filtration not working properly.
There is also an outbreak described from Norway where seagulls probably contaminated the
unprotected water reservoir (Y. Andersson, Pers Comm).
E. coli 0157:H7
E. coli is an enteric organism and comprises the majority of the normal flora of the gut. More
than 400 different serotypes of E. coli produce verocytotoxin, and most of these have been
linked to human illness (Molbak and Scheutz, 2004). E. coli 0157:H7 is the most widely
recognised verocytotoxin-producing E. coli (VTEC) serotype and is now recognised as an
important cause of food and water-borne illness in developed and some developing countries.
While 0157:H7 is the most commonly identified VTEC serotype in North America and the UK,
non-0157 VTEC are much more common in most continental European countries and Australia
(Molbak and Scheutz, 2004).
High incidence of VTEC infections has been reported from regions of Canada, Scotland, and
Argentina. In most European countries the annual incidence may range from 1-4 infections per
100,000 population. However, few laboratories screen for non-0157 VTEC, which remain
undetected. Human cases of VTEC infections tend to peak in the summer months, with highest
incidence in young children (Molbak and Scheutz, 2004).
Humans are thought to be the major reservoir, but livestock, such as cattle, sheep and, to a lesser
extent goats, pigs and chickens, are a major source of E. coli 0157 (WHO, 2004). In endemic
areas, such a the UK, E. coli 0157 may be present in up to half of the cattle herds, but this may
be an underestimate (Molbak and Scheutz, 2004). Zoonotic waterborne transmission may
therefore play a role in the spread of the agents.
Faecal shedding of E. coli 0157:H7 appears to be highest in young weaned cattle and during the
summer. It is thought that production practices such as feeding practice and crowding may
contribute to the emergence of E. coli 0157:H7 in cattle underestimate (Molbak and Scheutz,
2004).
In South Africa and Swaziland, in 1992, thousands of people were affected by bloody diarrhoea
and several fatalities occurred. Most cases were from men who drank surface water in the fields
and women and children who drank borehole water. E. coli 0157 was isolated from 14.3% of 42
23
samples of cattle dung and 18.4% of 76 randomly collected water samples. It was concluded that
cattle carcasses and dung washed into rivers and dams by heavy rains after a period of drought
contaminated the water (Isaacson et al. 1993).
The pathogens have been detected in a variety of water environments and outbreaks have been
reported, although due to the susceptibility of the organism to water treatment processes,
waterborne infection is relatively rare in industrialised countries (Chalmers et al. 2000).
However, although relatively rare, when waterborne outbreaks do occur the public health
consequences may be devastating. An outbreak of illness caused by E. coli 0157: H7 occurred in
the farming community of Walkerton, Ontario, Canada in May 2000. Seven people died and
2300 illnesses were reported. The drinking water supply was found to be contaminated by
rainwater runoff containing cattle excreta (O’Connor, 2002).
Dev et al. (1991) report an outbreak of E. coli 0157: H7 in Scotland in 1990. Because of the hot
weather during the summer, water levels in the water supply extraction points were low. As a
result of this water from two subsidiary reservoirs was used. However, one of the reservoirs was
fed from a source which may have been contaminated by cattle slurry.
A large outbreak of E. coli 0157 occurred in Fuerteventura, Canary Islands in March 1997
(Pebody et al. 1999). The cases occurred in four different hotels and it was established that three
of the four hotels were supplied with water from a private well.
It has been estimated that 1 to 4% of UK cattle herds are infected with E. coli 0157, however,
one study reported a regional incidence of 16% in cattle (Jones, 1999). Excretion by cattle may
persist for 2 to 4 months and appears to be seasonal with excretion highest in the spring and late
summer. This reflects the start of the peak in reported human cases. E. coli 0157 can survive in
cattle faeces up to 7 weeks, in non-aerated cattle manure for more than a year and in cattle slurry
less than 10 days (Jones, 1999).
It has often been difficult to establish whether the contamination of water with E. coli 0157:H7
is due to bovine or human origin, since both cattle and humans may shed E. coli 0157 and other
VTEC. However, it is known that ruminants may shed VTEC for longer periods of time than
humans (Molbak and Scheutz, 2004).
Infections have also been traced to the consumption of raw goat’s milk (Czech Republic;
Bielaszewska et al. 1997), raw cow’s milk cheese (Italy; Conedera et al. 2004), cheese (France;
Deschenes et al. 1996), or swimming in open lakes (The Netherlands, Finland; Cransberg et al.
1996).
Enteroviruses
Enteroviruses are one of the most common causes of human infections. They are ubiquitous,
enterically transmitted viruses that have been estimated to cause about 30 million infections in
the USA each year (WHO, 2004). The source of human pathogenic viruses in water is most
likely faeces from infected individuals independent of their disease status. Bird and animal
waste is thought unlikely to contain human enteroviruses (Percival et al. 2004) although Van der
Poel et al. (2001) found that husbandry animals were infected with viruses that are very similar
to human viruses. Coxsackievirus B5, has been closely linked with the virus that causes swine
vesicular disease and swine have been experimentally infected with Coxsackievirus B5 (Monlux
et al. 1975). Human infections with swine vesicular disease virus have occurred (Brown et al.
1976). To date, there is no documented evidence for humans to become infected from animal
viruses present in water. However, water is commonly contaminated with both human and
animal faeces and it is plausible that waterborne transmission of animal viruses and reassortant
viruses can occur. Lack of documentation of these events may be due to the difficulty in
detecting viruses in water (Cliver and Moe, 2004).
Many rivers in Europe have treated sewage discharged into them and therefore are likely to
contain enteroviruses. The vast majority of enteroviruses in controlled waters originate from undisinfected, continuous point source sewage discharge (Percival et al. 2004). Sediments from
freshwaters have been shown to be associated with enteroviruses (Lewis et al. 1986). These
sediments may be resuspended by rainfall, strong winds, tides and currents thereby releasing
their attached viruses back into the water column. Enteroviruses in sediments may persist for
24
long periods and so be capable of affecting counts for some time after discharging of sewage or
sludge may have stopped (Percival et al. 2004).
Norovirus
Noroviruses (genus Norovirus, family Caliciviridae) are a group of related, single-stranded
RNA, nonenveloped viruses. Noroviruses are considered the most common viral etiologic agent
of epidemic waterborne viral gastroenteritis (Brugha et al. 1999). Although outbreaks can occur
year-round, marked seasonal patterns of outbreaks have been observed (Lopman et al. 2003).
These patterns differ in the Northern and Southern Hemispheres. In the Northern Hemisphere,
gastroenteritis caused by Norovirus is most common in the winter and early spring, whereas in
the Southern Hemisphere, outbreaks are most frequent in the spring/summer (Marshall et al.
2003).
Primary infection with Norovirus results from the ingestion of faecally-contaminated water or
food (Estes et al. 2000; Parshionikar et al. 2003). Secondary, infection is by person-to-person
transmission, aerosolised vomits, formites, and infected food handlers. Low level transmission
can occur via contaminated drinking water supplies (Leclerc et al. 2002) when surface or
groundwater supplies are contaminated (Schaub and Oshiro, 2000).
Pathogen loads in sewage and manure
Overview
As infected livestock have considerable potential for contaminating aquatic environments,
agricultural practices are an important source of contamination especially from Cryptosporidium
oocysts, Giardia cysts, and Campylobacter (Carey et al. 2004; Lack, 1999; Monis and
Thompson, 2003).
Table 9 Examples of pathogens and indicator organisms commonly found in raw sewage. Source:
Adapted from Yates and Gerba, 1998
Pathogen or indicator1
Bacteria
Campylobacter spp.
Clostridium perfringens2
E. coli
Salmonella spp.
Shigella
Viruses, including enteroviruses
Polioviruses (vaccine)
Rotaviruses
Norovirus3
Coxsackievirus4
Parasitic protozoa
Cryptosporidium parvum oocysts
Entamoeba histolytica
Giardia lamblia cysts
Helminths
Ascaris spp.
Ancylostoma spp.
Trichuris spp.
Disease or role
No. per litre
Gastro-enteritis
Indicator organism
Indicator organism
Gastro-enteritis
Bacillary dysentery
37,000
6 × 105-8 × 105
107-108
20-80,000
10-10,000
Indicator
Diarrhoea, vomiting
Diorrhoea
vomiting
1,800-5,000,000
4,000-850,000
1.8 x 107 cDNA copies
0-5000
Diarrhoea
Amoebic dysentery
Diarrhoea
Ascariasis
Anaemia
Diarrhoea
1-390
4
125-200,000
5-110
6-190
10-40
25
1
Many important pathogens in sewage have yet to be adequately enumerated, such as adenoviruses,
norovirus/SRS viruses and Hepatitis A
2
From Long and Ashbolt, 1994
3
Frrom Laverick et al. 2004
4
From Rueedi and Cronin, 2005
As well as direct run-off into surface waters, animal waste is often collected in impoundments
from which the wastes may infiltrate groundwater. Runoff could also enter an aquifer through a
poorly sealed well casing. As it has proved difficult to quantify the contribution of various
sources of contamination, a first step in characterizing the risk of nonpoint source contamination
from pathogens of livestock origin is to determine the potential environmental loading based on
animal prevalence (Table 8) and faecal shedding intensity.
Large amounts of solid and liquid waste generated by domestic sources can also compromise the
quality of the body of water that receives the waste, and this is the predominant source of human
enteroviruses. This is a particular problem where large numbers of people live in close
proximity. Obviously the contamination loads of sewage will depend on the health status of the
population. Table 9 provides examples of pathogen loads typically found in raw sewage.
Injection wells used for domestic wastewater disposal are of particular concern to
groundwater quality if located close to and up gradient of drinking water wells. As urban
areas grow, there is an increase in rainwater runoff caused by the addition of paved
surfaces. Storm water drainage wells may be used to dispose of this additional runoff,
particularly if the area is not served by storm sewers or has limited sewer systems and
can serve as a conduit to groundwater for runoff from streets.
Cryptosporidium and Giardia
Calves can excrete up to 1010 Cryptosporidium oocysts per day (WHO, 2004). Preweaned
ruminants appear to be particularly susceptible to infection by Cryptosporidium. Some neonates
taken at birth and immediately delivered to clean rooms began excreting oocysts 3 days later,
suggesting susceptibility to an extremely low exposure dose or in utero transmission (Fayer,
2004). In manure, samples of broiler chickens neither Cryptosporidium oocysts nor Giardia cysts
were found (Medema et al. 2001). Concentrations of oocysts as high as 14,000 per litre for raw
sewage and 5800 per litre for surface water have been reported (WHO, 2004). Medema et al.
(2001) sampled effluents of cattle, pig and poultry slaughterhouses and found <0.13-2.5, <0.1710 and <0.11-0.66 Cryptosporidium oocysts per litre and <0.11-4, <0.17-14 and <0.2 Giardia
cysts per litre.
Studies in the Netherlands have shown Cryptosporidium is consistently present in municipal
wastewater (Table 10). The average geometric mean of Cryptosporidium in untreated sewage
water has been reported at 540/l at STP Kralingseveer and 4650/l at Amsterdam Westpoort
(Hoogenboezem et al. 2001). This study showed that the biological treatment of wastewater
removed Cryptosporidium with an average of 1.3 and 1.5 log units. The levels found in the
Netherlands during this study were reported to be broadly in line with the levels in sewage water
reported in the international literature.
Table 10 Cryptosporidium and Giardia oocysts in waste and surface waters. After Rose (1990).
Study
Untreated sewage
Madore et al. (1987)
Rose et al. (1986)
Medema et al. (2001)
Probable source of contamination
Agriculture wastewater
Human wastewater
(geometric mean)
Cryptosporidium oocysts/L Cryptosporidium
Giardia
oocysts/L
cysts/L
2904
1864
<0.61120(3.7)
700-10000
10-60000
26
Treated sewage
Ongerth and Stibbs
(1987)
Rose et al. (1988)
Rose et al. (1986)
Medema et al. (2001)
1.53
1.0*
1.09
0.58
0.25-11(0.35)
0-1350
0-700
*possible agricultural impact as well.
Despite apparent widespread infection of wild mammals with Cryptosporidium, data
documenting the extent of their contribution to pollution of surface waters are lacking (Fayer,
2004). Other studies have shown that by protecting a watershed with peripheral fencing results
in lower mean concentrations of Giardia and Cryptosporidium cysts (Ong et al. 1996).
LeChevallier et al. (1991) found fully protected watersheds had lower Giardia, but not
Cryptosporidium cyst concentrations in watersheds of limited access, compared with those with
recreational and agricultural activities, or those with sewage and industrial discharge.
Campylobacter and E.coli O157
Table 11 shows estimated quantities of slurry, poultry litter and farmyard manure, produced in
the UK per year, and the estimated accompanying loads of Campylobacters, from Stanfield and
Gale (2002). Table 12 shows corresponding values for E. coli 0157. The tables illustrate that
some animals are capable of shedding very high numbers of Campylobacters and E. coli 0157.
However, these figures do not tell us how much reaches the surface water and in those studies,
which have detected and quantified the pathogens of interest in source waters we have little
information on whether human infectious species were present.
Table 11 Manures produced in the UK per annum, and estimated Campylobacter content (Stanfield
and Gale, 2002).
Anima
l
Type of
waste
Cattle
Slurry
Pig
Farm yard
manure
Slurry
Sheep
Layers
Broiler
s
Farm yard
manure
Farm yard
manure
Farm yard
manure
Litter
Million
tonnes
pa
25
Prevalence of
campylobacter
(%)
60
28
60
3.3
80
6.7
80
1.3
75
1.3
50
1.6
80
Count/g
Mean
7.51
x104
6.51
x104
5.20
x104
5.30
x104
1.06
x104
1.95
x104
1.95
x104
Proportio
n of faecal
material
0.9
Campylobacte
r load per
annum
1.13 x1018
0.6
1.09 x 1018
0.9
1.37 x1017
0.6
2.84 x1017
0.6
1.03 x1015
0.9
1.27 x1016
0.5
2.50 x1016
Table 12. Manures produced in the UK per annum, and estimated E. coli 0157 content (Stanfield
and Gale, 2002).
Anima
l
Cattle
Type of
waste
Slurry
Million
tonnes
pa
25
Prevalence of
E. coli 0157(%)
Count/g
Mean
16
3.01x104
Proportio
n of faecal
material
0.9
E. coli 0157
load per
annum
1.08 x1017
27
Pig
Sheep
Layers
Broiler
s
Farm yard
manure
Slurry
Farm yard
manure
Farm yard
manure
Farm yard
manure
Litter
28
16
3.3
0.3
6.7
0.3
1.3
2
1.3
1
1.6
1
6.51
x104
3.75
x104
3.75
x105
2.45
x104
1.25
x104
1.25
x104
0.6
1.75 x 1017
0.9
3.34 x1014
0.6
4.52 x1015
0.6
3.82 x1014
0.9
1.46 x1014
0.5
1.20 x1014
Enterovirus and Norovirus
Enteroviruses are commonly found in sewage in relatively low numbers. Medema et al. (2001)
reports numbers between 34-190 L-1 in untreated sewage and 0.27-0.53L-1 in the effluent of two
sewage treatment plants in the Netherlands. Rueedi et al. (2004) found 0-5500 L-1 in raw
sewage from Doncaster (UK), where Coxsackievirus B2, B3 and B4 and Poliovirus type3 were
detected (Rueedi et al. 2004).
Noroviruses are found in sewage in low numbers. Schvoerer et al. (2000) reported noroviruses at
the entry and the exit of a sewage treatment plant in France. In Doncaster, UK, Noroviruses were
consistently found at different sewage sampling locations (Rueedi et al. 2004). Information
about norovirus numbers is generally rather rare because they cannot be cultured and the semiquantitative data come from RT-PCR on serial dilutions (see Section 2). However, there is not
much known about actual numbers of noroviruses in manure.
Summary
Exposure to waterborne pathogens in drinking water is a serious public health concern.
Therefore, it is important to determine the sources of pathogens in a watershed and to quantify
their environmental loadings. The natural variability of potentially pathogenic micro-organisms
in the environment from anthropogenic, natural, and livestock sources is large and is difficult to
quantify. As such it is impossible to rank the various sources and transmission routes in terms of
relative importance to human disease. More adequately, risks depend much on the specific case
and need to be considered in the local context. This is, of course, a big challenge for water
and/or health managers because no general recipe can be provided and, additionally,
requirements are changing. For this reason, it is important to determine different sources to
enable a distinction between their relative importance. Currently, much research is being
conducted on the development of methods to source pathogens, which will further increase our
ability to determine accurately the sources of inputs of pathogens into source waters.
28
Persistence of pathogens in the environment
The survival conditions for pathogens once voided from the animal organisms may be
unfavourable. Nevertheless, some can survive for extended periods - enteric bacteria for
example, have been shown to survive for at least up to sixty days (Fenlon et al. 2000), in what
are considered least hospitable environments such as on fabrics and plastics (Neely, 2000;
Robine et al. 2000). Antibiotic resistant strains of E. coli and Streptococcus faecalis were found
to persist in high numbers over a period of at least 32 days in saturated soil conditions
(Hagedorn et al. 1978).
Microbial activity in sediments is encouraged by the presence of organic matter (Ferguson,
1994; Millis, 1998). It is therefore possible that in nutrient-rich environments, micro-organisms
may survive in sediments for extended periods of time (Davies et al. 1995). When assessing
pathogens risks within a water body, it is important to determine whether pathogen resuspension
may occur within the time of survival of the pathogen. The resuspension of pathogens from
sediments due to turbulence at the benthic boundary, attributable to internal waves or wave
action of leeward shores, may present an unanticipated pathogen risk. The prolonged survival
and accumulation of micro-organisms in sediments, as well as the likelihood of their being
desorbed by dilution or water turbulence indicates that sediments, as well as surface waters,
should be assessed when estimating health risks (Geldrich, 1970; LaLiberte and Grimes, 1982;
Ferguson, 1994; Davies et al. 1995; Brookes et al. 2004).
Persistence of pathogens in surface waters
The persistence of pathogens in the aquatic environment is a function of both survival and
transport. Factors that control inactivation include temperature, salinity, pressure, solar radiation
and predation.
Temperature
Temperature is probably the most important and best investigated factor influencing the
inactivation of bacteria and viruses in the environment.
Table 13 Effect of temperature on inactivation of micro-organisms [days]
Organism
Temperature (oC)
10
20
30
Half life (days)
35-69
23
21
Cryptosporidiu
m
Giardia
2.2-4.6
3.6-7.7
Campylobacter
0.2-1.4
E. coli 0157
Enterovirus
5-100
1.6-69
0.080.16
5-100
0.24-14
Norovirus
3-30
0.05-5.8
39
Reference
Medema et al. 1997; Jenkins
et al. 2002
Mohammed et al. 2004;
deRegnier et al. 1989
Talibart et al. 2000; Blaser et
al. 1980; Lund, 1996
Nasser and Oman, 1999
Blanc and Nasser, 1996;
Hurst and Gerba, 1980;
Nasser et al. 2002; Nasser
and Oman, 1999; Yates et al.
1985
Gassilloud et al. 2003
Laboratory studies have demonstrated a negative correlation between water temperature and the
survival of coliform bacteria and enteric viruses, although the magnitude of the effect varies
29
between different strains. In general, these studies have measured survival times at three
temperatures: 10oC, 20oC and 30oC (Table 13). At the lower temperature enteric bacteria, such as
E. coli, have a half-life of several days, but at the higher temperature the half-life may be a short
as several hours. A similar effect has been observed for virus inactivation, although the survival
times are considerably longer than the survival times of bacteria. The influence of temperature
on the migration of bacteria and viruses is currently unknown.
Cryptosporidium
Several studies have looked at the effect of temperature on the infectivity and/or viability of
Cryptosporidium (Jenkins et al. 1997; Robertson et al. 1992 and Walker et al. 2001). The
general relationship between temperature, freezing time and infectivity is that C. parvum can
retain viability and infectivity after freezing and the oocysts can survive longer at higher
freezing temperature (Feyer and Nerad, 1996). Oocysts are not very good at surviving freezing
(Robertson & Gjerde, 2003) Anderson, (1985) found that extremes of temperature (above 60oC
and <-20oC) for 30 minutes will kill Cryptosporidium.
Giardia
Giardia cysts have been shown to be less resistant to environmental stress than Cryptosporidium
and have been shown to be viable for up to three months in cold raw water sources and in tap
water, with a range of 75-99% natural die-off (deReigner et al. 1989).
Bingham et al. (1979) also studied the effect of temperature on Giardia survival in water.
Storage at 8oC was found to result in the longest cyst survival (>77 days). Cysts stored at 21oC
retained viability for 5-24 days, whilst those kept at 37oC did not survive longer than four days.
Freezing and thawing of the cysts resulted in less than 1% viability, although this persisted for at
least two weeks.
Wickramanayake et al. (1985) determined that the optimum survival temperature for Giardia
cysts in water was 5oC.
Campylobacter
Little is known about the survival of Campylobacter except what is known from laboratory
experiments. These studies have shown that Campylobacter is only able to survive for a few
hours in adverse conditions with temperature being the major influencing factor. It was found
that the survival of Campylobacter was greater with decreasing temperatures, reaching several
days with 4oC. It was also found that survival was increased when Campylobacter was present
with other organisms within a biofilm (Buswell et al. 1998). Campylobacters have been shown
to survive in water for many weeks, and even months, at temperatures below 15oC (Buswell et
al. 1998; Holler et al. 1998). Buswell et al. (1998) showed mean survival times of
Campylobacter strains of 202, 176, 43 and 22 hours at 4, 10, 22 and 37oC respectively.
E. coli 0157
So far, research has shown that E. coli 0157 is no more persistent in the environment or resistant
to water-treatment processes than non-pathogenic E. coli (Percival et al. 2004). Table 12) Maule
(2000) showed that in soil cores containing rooted grass, VTEC 0157 can survive for 130 days at
18oC. VTEC 0157 could therefore enter water sources via surface runoff or drainage systems.
Once VTEC 0157 has entered freshwater sources it is able to survive for many days especially at
low temperatures. Rice et al. (1992) inoculated two human strains into well water at densities of
106 to 107 cfu per ml, and found no significant reductions in numbers after 7 days at 5ºC and
20ºC. A 3.5 log reduction was seen after 70 days at 5ºC. Chalmers et al. (2000) showed better
survival at 8ºC in reservoir water and recreational lake water compared with 15o and 25ºC. An
overall decrease of <2 log10 units was seen over a 13-week period at 8ºC.
E. coli 0157 has been shown to survive for up to 21 days in water but is as susceptible to
chlorination, temperature and source as any other E. coli strain. The lower infectious dose of E.
coli 0157 does potentially increase the risk of infection from biofilms in water but there have
been no outbreaks or studies of sporadic cases of E. coli 0157 implicating inadequately
disinfected water supply (Percival et al. 2004).
Flint (1987) examined the long-term survival of E. coli in river water. In sterile river water E.
coli was found to survive for up to 260 days, at temperatures ranging from 4-25oC, with no loss
of viability. According to Nasser and Oman (1999) the inactivation rate of E. coli was higher
30
than that of hepatitis A and poliovirus at lower temperatures, regardless of water type. Figure 1
shows a summary of inactivation rates of E. coli as function of temperature.
Lansbury and Ludlam (1997) and Armstrong et al. (1996) have identified Enterohemorrhagic E.
coli from 0.9 to 8.25% of healthy cattle in the UK. VTEC 0157 has been shown to survive for
long periods in bovine faeces, depending on temperature and moisture content. It is thought that
E. coli 0157:H7 may survive better in municipal water as compared with surface water and may
enter a viable but non-culturable state in both municipal and environmental water (Wang and
Doyle, 1998). This emphasises the importance of catchment management to ensure that numbers
of E. coli 0157:H7 in the source water remain low.
E.coli
inactivation rate [day-1]
10
1
0.1
0.01
0.001
0
5
10
15
20
25
30
35
Temperature [C]
Figure 1 Inactivation rates of E. coli as function of temperature (Nasser and Oman, 1999; Schijven
et al. 2000; Allwood et al. 2003; Gordon and Toze, 2003; Lee and Schiff, 2004).
Enteroviruses
Enteric viruses are renowned for their ability to survive for prolonged periods in aquatic
environments. Water temperature has been shown to be the dominant factor in determining
survival of viruses (Figure 2). Usually, increased temperature results in increased mortality
(Feacham et al. 1981; Lo et al. 1976; Ward et al. 1986; Olson et al. 2004b).
Studies conducted by Kutz and Gerba (1988) demonstrated the survival ability of enteroviruses
in freshwater sources. Mean inactivation rates were given as 0.325 log10 d-1 for polluted river
sources; 0.25 log10 d-1 for unpolluted river sources; 0.374 log10 d-1 for impounded water; 0.174
log10 per day for groundwater.
Hurst et al. (1989) showed that temperature affected the survival of coxsackievirus B3,
echovirus 7 and poliovirus 1 in samples of freshwater collected from five different sites. The
average amount of viral inactivation was 6.5-7.0 log10 units over 8 weeks at 22oC; 4-5 log10 units
over 12 weeks at 1oC, and 0.4-0.8 log10 units over 12 weeks at –20oC.
Although a number of studies have been conducted on the survival of viruses in soils in North
America, few studies have been conducted in Europe. Carrington et al. (1998) suggest that in
soil conditions of the UK, where mean soil temperatures do not tend to exceed 15oC at 10cm
depth in summer, and are about 5oC in winter, viral decay rates would be slow, with decimal
reduction times from 24 days to over 100 days. They also considered that cultivation of soil after
sludge application would encourage viral decay by encouraging evaporation.
Pathogen inactivation by low temperatures is only relevant where ice cover persists and would
therefore only affect those pathogens in the upper boundary. The formation of ice cover
contributes to stratification and so riverine intrusions would still move quickly through the
storage introducing fresh pathogens to the system and potentially resuspending previously
31
settled pathogens from sediments. Consequently, freezing of the water body does not necessarily
negate the pathogen risk. In addition, low water temperature may actually prolong the pathogen
survival. Sattar et al. (1999) showed that the rate of inactivation of Giardia at –20oC is faster
than inactivation of Cryptosporidium at the same temperature with a 1-log10 reduction in
viability in the first 12 hours and most Giardia cysts not viable after 24 hours. Freezing may not
have a similar impact on other pathogens such as E. coli.
Studies on norovirus indicate that the virus is fairly resistant to temperature and is not
inactivated by 30 minutes at 60oC (Keswick et al. 1985).
Enterovirus
inactivation rate [day-1]
100
10
Poliovirus
Echovirus
Coxsackievirus
1
0.1
0.01
0.001
0
10
20
30
Temperature
Figure 2 Inactivation rates different enteroviruses (O'Brien and Newman, 1977; Estes et al. 1979;
Hurst and Gerba, 1980; Yates et al. 1985; Jofre et al. 1986; Powelson and Gerba, 1994; Enriquez et
al. 1995; Blanc and Nasser, 1996; Nasser and Oman, 1999; Alvarez et al. 2000; Nasser et al. 2002;
Gordon and Toze, 2003; Schijven et al. 2003)
Salinity
Fayer et al. (1998) and Robertson et al. (1992) have shown that salinity effects the viability of
Cryptosporidium only at concentrations in excess of 20‰. Mansoury et al. (2004) showed that
infectivity and viability of Giardia and Cryptosporidium was affected at salinities of 50‰. As
potable water sources will generally have much lower salinity levels than this, it is unlikely that
salinity will have any effect on the inactivation of C. parvum in potable waters. However,
salinity is certainly an important factor in bathing waters. There is no reason to suggest that the
other pathogens of interest would be affected by the salinity levels of potable waters either.
Variations in salinity of surface water runoff may have an effect on pathogen mobility –
Bradford and Schijven (2001) found that dispersion decreased with increasing solution salinity.
Pressure
The effects of pressure on the inactivation of Cryptosporidium in drinking water sources are
negligible. Slifko et al. (2000) found that pressures in excess of 5.5·108 Pa were required to
make oocysts nonviable. This is equivalent to a water depth of 55,000 m. Therefore, it is
unlikely that the pressure exerted by water bodies used for drinking water would have any effect
on the other pathogens of interest.
pH
The effect of pH on the survival of pathogens in the environment has not been studied
extensively and the impact can only be inferred from laboratory investigations of the
physiological characteristics of the bacteria and the effect on the structural integrity of viruses.
32
In general, every species of bacteria has a narrow pH range that is optimum for growth.
Depending on the normal environment of the organism, the pH requirements can range from
highly acidic to highly alkaline: for many human pathogenic bacteria the optimum pH is close to
neutral. Despite having a preference for a narrow pH range, most species of bacteria can tolerate
a short exposure to a much broader range of pH. Outside these limits, the organisms are rapidly
killed. It is likely that pH affects the survival of viruses by altering the structure of the capsid
proteins and viral nucleic acid. However, it is unlikely that pH will be an issue at the ranges
found in surface water used for drinking water abstraction. It may affect migration of pathogens.
Solar radiation and inactivation of pathogens
Linden et al. (2001) showed that the most lethal wavelengths on the inactivation of
Cryptosporidium oocysts occur between 250 and 270 nm (UV-C). According to Brookes et al.
(2004) for summer conditions at mid-latitude, a typical measurement for incoming radiation
around midday is 1000 Wm-2, which, outside of the earths ozone layer, corresponds to
approximately 30 Wm-2 of UV light. For a one-hour exposure, this would correspond to a
cumulative UV light dose of 10800 mJcm-2. However, most of the UV-light is adsorbed by the
ozone layer allowing only a small proportion of approximately 100 mW/m-2 (or 36 mJ/cm-2 for 1
hour exposure) transmit to the earths surface of which the bulk part is UV-A light (315-400 nm).
In fact, the proportion of highly effective UV-C light is orders of magnitude smaller than that.
(Craik et al. 2001) investigated the effectiveness of low-pressure UV treatment facilities which
are normally in the range of 20-120 mJcm-2 and found inactivations of Cryptosporidium parvum
of more than 2 to 5 log units for intensities higher than 10 mJcm-2. For small intensities such as
under natural conditions in lakes and reservoirs it is very unlikey that UV light has the potential
to inactivate Cryptosporidium. Although repair of UV damaged DNA has been shown, oocysts
did not recover infectivity (Shin et al. 2001).
Giardia cysts have been shown to be sensitive to UV light at low doses under laboratory
experimental conditions. A 2-log inactivation was seen at a dose of 3 mJ/cm2 (Mofidi et al.
2002) and 3-log inactivation at 20-40 mJ/cm2 (Campbell and Wallis, 2002), both studies using
an animal model to assess viability/infectivity.
By contrast, Obiri-Danso et al. (2001) showed that natural populations of C. jejuni and C. lari
exposed to artificial sunlight became non-culturable within 30 minutes, with T90s of 25 minutes
and 15 minutes respectively.
The effect of sunlight on enteroviruses was studied by Gerba et al. (2002). 3-log reductions of
echovirus 1, echovirus 11, coxsackievirus B3, coxsackievirus B5, poliovirus 1, and human
adenovirus type 2 were effected by doses of 25, 20.5, 24.5, 27, 23, and 119 mW/cm2,
respectively. Human adenovirus type 2 is the most UV light-resistant enteric virus reported to
date.
Light intensity has been identified as one of the most influential factor causing die-off of
coliforms in freshwater (Garvey et al. 1998). In illuminated water Barcina et al. (1989) found
that most E. coli cells were in a somnicell state after 72 hours. The rate of decrease in numbers
of culturable bacteria was found to be faster in seawater exposed to natural sunlight (400-700
nm) than in freshwater.
The lack of culture methods for norovirus means that a cultivable model virus is used to
investigate their inactivation. Duzier et al. (2004) used enteric canine calicivirus no. 48 (CaCV)
and the respiratory feline calicivirus F9 (FeCV) and correlated inactivation to reduction in PCR
units of FeCV, CaCV and a norovirus. Inactivation of suspended viruses was temperature- and
time-dependent in the range from 0-100oC. UV-B radiation from 0-150 mJ/cm2 caused dosedependent inactivation, with a 3 log10 reduction in infectivity at 34mJ/cm2 for both viruses.
Norovirus was never more sensitive than the animal caliciviruses.
Ammonia
Pathogen survival in manure heaps and slurry stores is adversely affected by pH, temperature
and ammonia generated (Bukhari et al. 1999; Jenkins et al. 1999). Cryptosporidium survival of
33
up to 176 days in drinking water or river water with an inactivation of 89-99% of oocysts has
been reported by Robertson et al. (1992).
Jenkins et al. (1999) has shown that free ammonia (NH3) has been found to cause high levels of
inactivation of Cryptosporidium oocysts at concentrations above 7 mmol/L. Increased levels of
ammonia are often seen in the hypolimnion of lakes towards the end of the stratified period, as
low dissolved oxygen concentrations result in sediment release of ammonia. These values are
typically much lower than 1 mg/L, and so the effect of free ammonia on Cryptosporidium oocyst
viability will be negligible in drinking water reservoirs (Brookes et al. 2004).
Predation of pathogens
The grazing of pathogens by aquatic invertebrates has a number of implications for transport
through a water body, settling characteristics and possible transmission to humans. The main
finding of laboratory studies is that microbial activity in the soil and groundwater reduces the
survival time of enteric bacteria and viruses. Evaluating the role of microbial activity normally
involves a comparison of survival times in sterile and non-sterile environments. There are
several conflicting reports regarding the influence of indigenous populations of micro-organisms
on the survival of enteric bacteria and viruses, ranging from increasing the rate of inactivation
through to having no effect to decreasing the rate of inactivation. For examples, the level of
heterotrophic bacteria in natural waters was found to influence the survival of Cryptosporidium
oocysts (Heisz et al. 1997).
Studies with thermotolerant coliforms as a whole, and E. coli in particular, have shown that the
concentration of the test organism can increase rapidly in sterile environments, but remains
static, or is reduced in non-sterile environments (Gerba and McLeod, 1976). If pathogens are
grazed by free-living protozoa or zooplankton and then excreted intact in a faecal pellet this may
change the settling behaviour of the pathogen. McCambridge and McMeekin (1980) showed that
predacious protozoa exerted a major influence on E. coli destruction during the first 2 days of a
10-day decline. Brown and Baker (1999) have summarised current knowledge about the
survival of bacteria in protozoa. In this review the authors point out that, apart from their
interaction with Legionella pneumophila and related species, the role of protozoa as a reservoir
for the maintenance of pathogenic bacteria in the environment, and as a possible vector for the
transmission of human and animal disease, has received little attention. However, there is some
evidence to suggest that the association between bacteria and protozoa represents an important
survival mechanism for some species of bacteria and a means of maintaining, or increasing
virulence of the organism. The full implication to public health and waterborne disease of the
association between bacteria and protozoa remains to be elucidated.
Hurst et al. (1980a) showed that the inactivation rate of two strains of enterovirus was more
rapid in non-sterile, aerobic environments than in sterile environments. By contrast, Matthess et
al. (1988) found no significant difference between the inactivation rates of several viruses in
sterile and non-sterile groundwater. Gordon and Toze (2003) looked at the effects of
temperature, oxygen, nutrient levels and presence/absence of micro-organisms on survival of
poliovirus and coxsackievirus. They concluded that the latter was the most influential factor in
virus survival in groundwater.
A summary of the major influencing factors on pathogen survivals are listed in Table 14.
Table 14 Summary table with correlation trend between parameter and pathogen inactivation rate
in brackets
Cryptosporidium
Giardia
Campylobacter
Solar
radiatio
n
Temperatur
e
Salinity
Pressure
Predation Ammonia
Medium
(+)
Medium
(+)
High (+)
High (+)
Medium (+)
Low (+)
Medium (+)
Medium (+)
High (+)
Medium (+)
Low (+)
Medium (+)
Medium (+)
High (+)
Medium (+)
Low (+)
Medium
Medium (+)
34
E. coli
High (+)
High (none)
Medium (+)
Low (+)
Enteroviruses
High (+)
High (+)
Medium (+)
Low (+)
Norovirus
Likely
High (+)
Likely High (+)
Unknown –
likely
Medium (+)
Low(+)
(+/-)
Medium
(+/-)
Medium
(+/-)
Unknown,
likely
Medium (+)
Medium (+)
Medium (+)
Unknown
35
Review of knowledge on transport of pathogens
Transport in surface water
Once micro-organisms are released into the environment, their fate becomes subject to various
aspects of the environment, some of which have already been discussed in section 4. The first
consequence is that organisms become dispersed as they are added to water or soil. Microorganisms can be transported through the environment along with the soil, either as wind carried
particles or as material that is carried down slope in landslide. The movement of released
organisms is greatly facilitated by the flow of water, which can occur either on the surface or in
the subsurface. Riverine inflow is considered to be a major source of pathogens. The behaviour
of these inflows is therefore important. Warm inflows will flow over the surface of a lake as a
buoyant surface flow and cold, dense inflows will sink beneath the lake water where they will
flow along the bathymetry towards the deepest point. In both situations, the inflow will entrain
water from the lake, increasing its volume, changing its density and diluting the concentrations
of pathogens and other properties (Brookes et al. 2004).
The deepest point of most drinking water reservoirs is where the off-take is located. Where
there is a dense underflow, and the density of the underflow matches that of the adjacent lake the
underflow will become an intrusion. In some cases, the underflow is denser than any water in the
lake and it will flow to the deepest point. As oocysts of Cryptosporidium for example survive
longest in cold and dark water, this means that this situation will present the greatest risk to the
quality of the water.
There are a number of factors which are of importance in the determination of the hydrodynamic
distribution of pathogens in lakes and reservoirs. These include: the speed at which an inflow
travels through a lake, its entrainment of lake water and resulting dilution of its characteristics
and its insertion depth (Brookes et al. 2004).
Settling
The aggregation of pathogens to particulate material, or the integration of pathogens within a
matrix of organic material, will influence the rate of pathogen settling. In addition, this rate may
also be affected by predation of pathogens and the subsequent incorporation of pathogens into
faecal pellets. The surface charge of the particles is important in the interaction between the
particles (Ongerth and Pecoraro, 1996). Drozd and Schwartzbrod (1996) suggest that
aggregation of Cryptosporidium oocysts to particles and to each other is pH-dependent which is
a result of the pH adjusting the hydrophobic and electrostatic nature of the oocyst surface.
Studies have shown that there is little variation over a small range of pH values as found in
drinking water reservoirs. Drozd and Schwartzbrod (1996) were unable to show a relationship
between pH and the size of the aggregated particles. The size of the particles with which
Cryptosporidium for example, is associated with is a major factor influencing the transport of
these pathogens across a landscape, river or reservoir. If Cryptosporidium is associated with
large particles, there is a greater chance of interception, or settling, and so less of a risk than if
they are associated with small particles (e.g. clay) or transported as single unattached oocysts
(Brookes et al. 2004).
There have been few studies published on the dispersion, survival and viability of pathogens
excreted in faecal matrices (Jenkins et al. 1999; Bradford and Schijven, 2002) but it is possible
that the mastication of plant material by cattle and the subsequent scouring of the stomach wall,
which dislodges oocysts, will have a significant impact on the interaction between
Cryptosporidium and particles (Brookes et al. 2004).
Aggregation is another factor affecting the size of pathogen-associated particles. Ongerth and
Pecoraro (1996) indicate that Cryptosporidium oocysts are strongly negatively charged at neutral
pH. Consequently, they may be adequately aggregated and flocculated during conventional
water treatment but may not adsorb well on natural clays in the environment. Dai and Boll
(1993) suggested that oocysts do not attach to natural soil particles and would travel freely in the
36
water. This theory has been supported by Considine et al. (2000; 2001) but they also concluded
that protein-linked tethering between silica and oocysts can occur and may facilitate adhesion.
Since this interaction relies on contact, there must be adequate turbulence in the system to
increase the probability of collision between particles and oocysts.
There appears to be two conflicting arguments as to whether Cryptosporidium is associated with
particles. The surface charge of oocysts suggest that they would not absorb readily to particles,
but the very high settling velocities recorded by Hawkins et al. (2000) and Medema et al. (1998)
suggests that, at least in certain situations, oocysts must be associated with larger particles. One
alternative is that the oocysts may be physically mixed within an organic matrix of faecal
material and/or soil particles during entrainment in surface water runoff (Brookes et al. 2004).
Feng et al. (2003) showed that suspended particles present in reservoir water contributed to
enhanced recovery of Cryptosporidium parvum oocysts and that particle size and concentration
could affect oocyst recovery. The optimal particle size was found to be in the range of 5-40 µm,
and the optimal concentration of suspended particles was 1.42 g for 10 litres of tap water.
Viruses such as coxsackievirus type B3 appear to readily adsorb to sediment. It has been shown
that greater than 99% of these viruses adsorption to sediment (La Belle and Gerba, 1979).
Gantzer et al. (2001) also showed significant adsorption to soil of somatic coliphages, F-specific
RNA phages and faecal coliforms from wastewater (61%, 78% and 86% respectively).
Since pathogens remain viable in the sediments of a lake or reservoir for variable lengths of
time, it is important to consider the importance of their resuspension and subsequent
redistribution. Sediment resuspension occurs when turbulent velocity fluctuations reach a
critical level (Brookes et al. 2004).
Transport in sediments
The transport of pathogens through the water body and through sediment is another factor which
will determine the contamination level of the water body. Concentration of Giardia oocysts, for
example has been shown to be positively correlated to water flow and turbidity levels (Atherholt
et al. 1998). Although little is known about the mechanisms that determine the transport of
oocysts in river water, it is likely that transport by the flow of the water is the major determinant
(Medema and Schijven, 2001). Zuckerman et al. (1997) demonstrated degradation of
Cryptosporidium oocysts in the presence of Serratia marcescens, a bacteria with high chitinolytic
activity.
Contradictory results on the effect of organic matter on virus behaviour have been reported in
the literature. Gerba (1984) showed that the presence of organic matter can reduce virus
attachment and thus facilitate virus transport by providing additional negative charges, covering
positively charged sites, or competing with viruses for attachment sites. On the other hand, Bales
et al. (1995) and Kinoshita et al. (1993) showed that organic matter inhibits virus transport by
promoting hydrophobic interactions between viruses and grain surfaces.
Different classes of organisms have specific characteristics, such as size or charge, which
determine their movement and survival in the aquatic environment and their susceptibility to
various water and wastewater treatment processes (see section 5). Knowledge of these
characteristics can help in the design of effective barriers or control strategies.
Transport in the subsurface
Introduction
Some, perhaps many instances of groundwater receptor contamination will occur by rapid
transport pathways accidentally introduced by human intervention and connecting the
contamination source to the groundwater abstraction point. Such pathways could include, for
example, inadequate sanitary completion of springs, wells and boreholes, or the presence of a
forgotten conduit connecting the source of contamination to the groundwater abstraction point.
The implementation of management actions to reduce faecal contamination close to the
abstraction point, or the rehabilitation or improvement of the well or spring is usually sufficient
to control access of pathogens to the water source.
37
Rapid transport pathways cannot, however, explain all groundwater source contamination events
and it is now widely accepted that the transport of microbial pathogens within groundwater
systems is a significant mechanism for waterborne disease transmission. This section deals with
the factors that control the transport and attenuation of pathogens into and through groundwater.
As discussed, the presence of pathogens in water is due to a number of factors, controlling input,
survival and transport, depending on the type of water and on aquifer characteristics in the case
of groundwater. Rain is an important vector in the presence of pathogens in water, and
temperature and the presence of cattle, and other livestock are particularly significant factors for
surface water and groundwater respectively (Schaffter and Parriaux, 2002). In groundwater, the
duration of a contamination event is a function of hydrodynamic properties. Therefore, a spring
with a high dilution rate for direct infiltration water near the spring is subject to negligible
contamination that disappears quickly. Conversely, a spring collecting high quantities of direct
infiltration water, will show high levels of contamination lasting several months after the cattle
or other potential source of contamination have left the catchment area. In extreme cases, the
persistence of contamination is so great that the presence of the source of contamination is no
longer significant at all. There is therefore the possibility that certain types of bacteria may
persist for considerable periods in a natural environment, probably adsorbed on soil particles or
on silts so as to increase their chances of survival (Schaffter and Parriaux, 2002).
The processes of dispersion, dilution, horizontal and vertical transport determines the
distribution of pathogens in the subsurface. Settling of pathogens particles works in conjunction
with these hydrodynamic processes.
Time scales of groundwater transport
The great variation in residence times in freshwater bodies is illustrated in Figure 3, which
emphasises the generally slow movement and long residence time of groundwater compared to
surface waters.
Figure 3 Water residence time in inland freshwater bodies (after Meybeck et al. 1989)
The short residence times within karstic and alluvial aquifers originate from the preferential flow
paths available for rapid groundwater flow. These flow paths are responsible for the usually high
contamination risks of these aquifers. A groundwater volume withdrawn at a certain location
always comprises water with a range of different ages. Old and well-filtered water is not likely
to be a source of contamination but even a very small amount of very young water can lead to
severe consequences, particularly for pathogens with high infectivities, such as those of interest
in this review. However, it is very difficult to determine such small quantities of young water
with independent methods. This leads to the limitation that the potential problems sometimes
cannot be detected before the first incident occurs. Therefore, protection of drinking water wells
38
will always underlay some sort of assumption considering the fastest pathways and their
probabilities.
Groundwater occurrence and storage
Some groundwater occurs in most geological formations because nearly all rocks in the
uppermost part of the Earth’s crust, of whatever type, origin or age, possess openings called
pores or voids. Geologists traditionally subdivide rock formations into three classes, according
to their origins and methods of formation:
•
Sedimentary rocks are formed by deposition of material, usually under water from
lakes, rivers and the sea, and more rarely from the wind. In unconsolidated, granular materials
such as sands and gravels, the voids are the spaces between the grains (Figure 4a). These may
become consolidated physically by compaction and chemically by cementation (Figure 4d) to
form typical sedimentary rocks such as sandstones, limestones and shales, with much reduced
voids between the grains.
•
Igneous rocks have been formed from molten geological material rising from great
depth and cooling to form crystalline rocks either below the ground or at the land surface. The
former include rocks such as granites and many volcanic lavas such as basalts. The latter are
associated with various types of volcanic eruptions and include lavas and hot ashes. Most
igneous rocks are strongly consolidated and, being crystalline, usually have few voids between
the grains.
•
Metamorphic rocks have been formed by deep burial, compaction, melting and
alteration or re-crystallisation of other rocks during periods of intense geological activity.
Metamorphic rocks include gneisses and slates and are also normally consolidated and with few
void spaces in the matrix between the grains.
Figure 4 Rock texture and porosity of typical aquifer materials (based on Todd, 1980). a) wellsorted, unconsolidated sediment with high porosity (e.g. alluvial sands); b) poorly sorted sediment
with low porosity; c) well-sorted sediment of porous pebbles; d) sediment whose porosity has been
diminished by deposition of mineral matter; e) rock with porosity increased by solution (e.g.
limestone); and f) rock with porosity increased by fracturing (e.g. granite)
39
In the more consolidated rocks, such as lavas, gneisses and granites the only void spaces may be
fractures resulting from cooling or stresses due to movement of the earth’s crust in the form of
folding and faulting. These fractures may be completely closed or have very small and not very
extensive or interconnected openings of relatively narrow aperture (Figure 4f). Weathering and
decomposition of igneous and metamorphic rocks may significantly increase the void spaces in
both matrix and fractures. Fractures may be enlarged into open fissures as a result of solution by
the passing groundwater (Figure 4e). Limestones are largely made up of calcium carbonate and
therefore particularly susceptible to active solution, which can produce the caverns, swallow
holes and other characteristic features of karstic aquifers.
Transport mechanisms
The processes of pollutant attenuation in groundwater systems can be subdivided into dilution,
retardation and elimination or transformation. Solutes or particles are carried in moving
groundwater by advective transport, which is usually modified by the processes described
below, causing mixing and dilution with uncontaminated water on the flow pathway through the
aquifer. Advection is the transport of non-reactive solutes at the same speed and in the same
direction as the average water velocity. This concept is applicable for both saturated and
unsaturated contaminant transport with the only difference that transport through the unsaturated
zone is predominantly vertical whereas groundwater transport is predominantly horizontal.
Advection
Principally, there are two kinds of advective transport mechanisms observed in different rock
types: a) preferential flow and b) matrix flow. Groundwater flow in granites or limestones is
dominated by preferential flow where flow velocities and directions are often unknown, difficult
to assess and highly variable. In sandstones and alluvial sands or gravels groundwater flow is
dominated by matrix flow where groundwater is supposed flow regularly and relatively slow as
a homogeneous package. In shales, finally, both processes are similarly important.
The matrix-flow of groundwater though an aquifer is governed by Darcy’s Law, which
states that the rate of flow is directly proportional to the hydraulic gradient:
Q = K ⋅i ⋅ A
[1]
where Q is the rate of flow through unit area A under hydraulic gradient i. The hydraulic
gradient (dh/dl) is the difference between the levels of the potentiometric surface at any two
points divided by the horizontal distance between them. The parameter K is known as the
hydraulic conductivity, and is a measure of the permeability of the material through which the
water is flowing. For clean, granular materials, hydraulic conductivity increases with grain size.
Typical ranges of hydraulic conductivity for the main types of geological materials are shown in
Figure 5.
40
Figure 5 Range of hydraulic conductivity values for geological materials (based on Driscoll, 1986
and Todd, 1980)
Dispersion
When a small volume of a solute, for example a pollutant or a tracer, is released into
groundwater it will spread from the advective flow path to form a plume of diluted solvent. The
plume broadens both along and perpendicular to the groundwater flow direction (Figure 6). The
processes that contribute to this spreading or hydrodynamic dispersion are molecular diffusion
and mechanical dispersion. Molecular diffusion is the movement of solute ions in the direction
of the concentration gradient from high towards low concentrations. This movement originates
from the thermal-kinetic energy of the solute ions but is a very slow process, which is only
important at low groundwater velocities. In consolidated aquifers such as sandstones and
limestones (Figure 4e and f), diffusion may occur between waters of different concentrations in
the fractures and the matrix. Mechanical dispersion arises from (a) the tortuosity of the pore
channels in a granular aquifer and of the fractures in a consolidated aquifer and (b) the different
speeds of groundwater in flow channels of different widths.
The first produces lateral dispersion perpendicular to the flow direction and the second
longitudinal dispersion in the direction of flow. The latter is generally much stronger than the
former, as shown in Figure 6.
41
Figure 6 Dispersion in a homogeneous isotropic aquifer. A fixed volume of tracer is released at the
injection point A at time 0. At time t the tracer has reached B; after time t’ it has reached C; after
time t’’ it has reached D (after Price, 1996)
Dilution is, therefore, most effective below the water table but can have some impact in the
unsaturated zone. Concentrations of surface-derived pollutants such as nitrate from agricultural
fertilisers may be related to recharge volumes. In the UK for example, higher nitrate
concentrations (in mg/l) are observed in the lower recharge of the drier east and south of the
country than in the west, even though the same amounts (in kg/ha) may be leached. Dilution can
also be a factor in reducing pollutant concentrations (but enhancing pollutant transport) where
additional recharge comes from leaking water distribution systems, urban drainage or excessive
irrigation.
Pathogen attenuation: filtration and sorption
The processes of retardation include filtration and sorption as a result of the geological
environment. Filtration is a process that affects particulate contaminants (e. g. organic/inorganic
colloids or microbes) rather than solutes. Particles larger than pore throats diameters or fracture
apertures are prevented from moving by advection and are therefore attenuated within the soil or
rock. Particle sizes of selected pathogens are listed in Table 15 and shown in Figure 7.
Table 15 Sizes of selected pathogens
Pathogen
Cryptosporidiu
m parvum
Giardia
Campylobacter
length width
[µm]
[µm]
4.5-5.4 4.2-5.2
12-15
0.2-0.5
0.5-5
6-8
E.coli O157
2-6
1.1-1.5
Enterovirus
0.0220.03
Norovirus
0.027-
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42
0.03
10-3m
1 mm
10-6m
1 µm
protozoa
10-9m
1 nm
bacteria
PATHOGEN
DIAMETER
viruses
fissures
apertures
sands
sandstone
pores
1Å
FISSURE
APERTURE/
PORE SIZES
limestone
chalk pores
silt pores
Cryptosporidium
oocyst
© NERC. All rights reserved.
Figure 7 Pathogen diameters compared to aquifer matrix diameters
Sorption is a process by which organisms become attached to particles of clay or organic matter
in the soil or aquifer to remove them from the water. The sorption process depends on the nature
of the charges on the micro-organism and the particles, and therefore on the pH of the water. If
the pH of the water changes, the distribution of surface charges changes and adsorbed particles
may be released from their sorption sites and reintroduced into the groundwater. This reverse of
the process is known as desorption.
Micro-organisms can be seen as charged colloid particles moving with the water. All colloids
suspended in water have electrostatic potentials associated with their water-cell, boundary
interface. The electrical charge associated with this surface is controlled by complex
physicochemical interactions between the microbe and the surrounding solution. Solution
chemical properties such as pH and ionic strength play an important role. At near-neutral pH
conditions (pH near 7) and typical natural groundwater ionic strengths, most suspended
microbes have a net negative surface charge. Most small-grained mineral surfaces (clays, silts,
and sands) in aquifers also usually have net negative surface electrostatic charges. Therefore, in
most groundwater environments (at pH ≈ 7), microbes are slightly repelled electrostatically by
mineral surfaces and thus tend not to adsorb. However, in some conditions of pH, mineral
characteristics, and other water chemistry factors, attractive electrostatic potentials are present
between mineral surfaces and microbes, and the microbe will tend to adsorb. One measure of an
electrostatic surface property of a microbe is its isoelectric point or point of zero net charge - the
pH at which its surface changes from a negative to a positive net charge. This isoelectric point
(pI) can be measured in the laboratory (Table 16).
Table 16 Isoelectric points (pI) of different pathogens.
Pathogen
pI
References
Cryptosporidium
3.9, 2.18,
2.37, 3.3
3.6, 2.2
4.8-4.9
4-5
Ongerth and Pecoraro, 1996; Drozd &
Schwartzbrod, 1996; Brush et al. 1998; Hsu and
Ongerth and Pecoraro, 1996; Hsu and Huang, 2002
Glenn-Calvo et al. 1994; Janvier et al. 1998
Lytle et al. 1999
Giardia
Campylobacter
E.coli O157
43
Enterovirus
Norovirus
Polio: 4 Gerba, 1984; Wossener et al. 2001
and 7.5
5.0
Goodridge et al. 2004
A summary of different approaches developed to model colloid transport are summarised in
Ryan and Elimelech (1996). Principally, two kinds of approaches can be distinguished, namely
chemical and mechanical. Chemical methods include the DLVO theory developed by (Derjaguin
and Landau, 1941) and (Vervey and Overbeek, 1948) which principally tries to explain colloid
sorption based on attractive and repulsive forces acting between collectors (aquifer matrix) and
colloids (virus, bacteria or protozoa). These forces are affected by changes in groundwater
chemistry, such as changes in pH, Eh or electric conductivity (see below). Mechanical
approaches try to estimate colloid filtration or removal using empiric solutions, involving flow
velocities, attraction/repulsion forces and diffusion (Rajagopalan and Tien, 1976; Ruckenstein
and Prieve, 1973; Spielman and Friedlander, 1974; Yao et al. 1971).
Table 17 summarises sorption rates of the different pathogens originating from column
experiments on sand columns. Note that these rates do not distinguish between sorption and
filtration and often, desorption is not considered.
Table 17 Sorption and desorption rates for pathogens in sand columns [day-1].
Cryptosporidium
Giardia
Campylobacter
E.coli O157
Enterovirus
Norovirus*
sorption
[day-1]
40-700
1.5-2 times
Cryptosporidiu
m
Between E.coli
and
Cryptosporidiu
0.1-1.5
desorption
[day-1]
References
Harter et al. 2000
Dai et al. 2004
Hijnen et al. 2004
Gagliardi and Karns, 2000;
Powelson and Mills, 2001
0.7-2.1
0.009-0.06 Schijven et al. 2003; Sobsey et
al. 1995
Not quantitative with current methods
Summary of major factors influencing pathogen transport
This section will only discuss the major factors affecting pathogen adsorption and desorption on
soil or rock material. Pathogen inactivation, which is an important factor for their transport, was
discussed in Section 0 The mechanisms by which microbial contaminants may undergo transport
and attenuation in the saturated and unsaturated zones have been described earlier in Section 0.
There now follows a description of the factors that control the degree of the impact of these
mechanisms. The potential for pathogens in faeces and wastewater to contaminate the
underlying groundwater is dependent on a number of factors including the physical
characteristics of the site (e.g. soil texture), the hydraulic conditions (e.g. wastewater application
rate, wetting/drying cycles), the environmental conditions (e.g. rainfall, temperature) at the site,
and the characteristics of the specific pathogens present in the water. The factors that influence
the transport and attenuation of pathogens in the subsurface have been the subject of a number of
reviews summarised by Bitton and Harvey (1992); Robertson and Edberg (1997); Schijven and
Hassanizadeh (2000); Vaughn et al. (1983); Yates and Yates (1988); Yates et al. (1985). Some
of the major factors influencing pathogen transport and attenuation are described in more detail
below.
44
pH
As mentioned in section 4, the most important factor controlling adsorption of micro-organisms
is the pH of the groundwater. Some authors have suggested that pH indirectly influences
pathogen survival by controlling adsorption to soil particles and the aquifer matrix. It is the
adsorption to surfaces that ultimately increases the survival time of the pathogens. Generally,
bacteria and viruses have negative surface charges generated by the level of ionisation of the
carboxyl and amine groups that are a major component of surface proteins. As the pH of the
medium changes, the ionisation of the two groups will change, causing a shift in the net strength
and polarity of the surface charge. At a specific pH, which is determined by the molecular
structure of the protein, the net charge will be zero; this is termed the isoelectric point of the
molecule. The isoelectric point has been determined for many different proteins and for a
number of virus strains. At pH values below the isoelectric point a virus will have a net positive
charge, whereas the charge will be negative at pH values above the isoelectric point.
Within the pH range of most unpolluted groundwater both the matrix surfaces and the surfaces
of the micro-organisms carry a net negative charge. Under these conditions the micro-organisms
will be repelled by most mineral grain surfaces. At low pH values the surface charge on the
micro-organisms will shift to being net positive, which will favour their adsorption to soils and
the aquifer matrix by electrostatic attraction. This hypothesis has been confirmed by several
groups of workers for both bacteria and viruses (Bitton and Harvey, 1992; Gerba and Bitton,
1984; Sobsey, 1983).
There are many complicating factors that can interfere with the mechanism discussed above.
One is that a given virus may have more than one isoelectric point and the factors responsible for
passage from one form to another are unknown at this time. Other factors, such as cations and
humic and fulvic acids, may also influence the net surface charge of the organism.
Whilst changes in pH may affect the mobility of micro-organisms in the subsurface, the
significance of this factor in any particular aquifer is uncertain. Robertson and Edberg (1997)
have noted that the pH of most unpolluted groundwater is generally very stable, and in their
experience falls within the near-neutral range of 6.5 to 8.5. This is because most exploited
aquifers are of sedimentary origin and therefore contain at least some calcite, which buffers the
groundwater pH by dissolution or precipitation respectively. There are exceptions, where the
buffering capacity of the aquifer is low (e.g. gneisses and granites) the pH is likely to be much
lower, frequently in the range 5.5 to 6.5. Robertson and Edberg conclude that it is unlikely that
significant changes in microbe mobility will occur due to these minor pH changes. This
assumption may be valid for many, stable groundwater systems, but in groundwater, and indeed
surface waters that are exposed to contamination from a variety of sources, which may be of
unknown and variable quality, for example sewage, pH may emerge as a dominant factor in the
mobility of pathogens.
Soil moisture content
Although some investigators have observed no difference between the inactivation rates of
viruses in dried and saturated soils (Lefler and Kott, 1974) the majority of reports have shown
that soil moisture content influences the survival of viruses in the subsurface. For example,
Hurst et al. (1980a) found that moisture content affected the survival of poliovirus in loamy
sand. The inactivation rate of poliovirus decreased as the moisture content increased from 5 to
15 per cent. However, further increases in soil moisture content reduced the survival time of the
virus. It was noted that the inactivation rate peaked near the saturation moisture content of the
soil (15 to 25 per cent), and was slowest at the lowest moisture contents (5 to 15 per cent). Soil
moisture has been reported to influence the fate of bacterial contaminants (Robertson and
Edberg, 1997), but the magnitude of the effect, and the value of any correlation has not been
described.
Salt species
45
The adsorption of micro-organisms onto surfaces in the groundwater system has been shown to
have two counteracting effects: It reduces the dispersal of the organism in the subsurface, but
increases the survival time of the organism in the affected area. If the prevailing geochemical
conditions in the groundwater system create opposing charges on the surface of the organism
and the aquifer matrix, adsorption will occur by electrostatic attraction. Frequently, however,
these conditions do not exist and the organism and the aquifer matrix each have a negative
charge.
The types and concentrations of salts in the environment can have a profound influence on the
extent of pathogen transport in the subsurface. Cations (positively charged inorganic species), in
particular multivalent cations such as Magnesium (Mg2+) and Calcium (Ca2+) can form a bridge
between the solid surface and the organism and significantly enhance adsorption. Clearly, the
concentration of the salt is also important, as this will influence the number of sites that are
available for binding as well as the number of bridges that can be formed between the two
surfaces. Thus the capacity for binding and the strength of the bonds will be affected by the salt
concentration. Several studies of virus and bacterial transport through simulated groundwater
systems have confirmed this hypothesis (Bitton and Harvey, 1992; Simoni et al. 2000; Sobsey,
1983; Taylor et al. 1981).
A decrease in the salt concentration or ionic strength of the soil water, such as would occur
during a rainfall event, can cause desorption of viruses and bacteria from soil particles (Gerba
and Bitton, 1984). This phenomenon has been observed in both laboratory and field experiments
(Landry et al. 1980; Wellings et al. 1975). Furthermore, there is evidence to suggest that only
small changes in the salt concentration can dramatically affect the mobilisation of some
organisms in groundwater systems (Redman et al. 1999). The implication of this discussion is
that salt concentration in the groundwater system may be of greater significance to pathogen
transport than pH, although it is important to consider that neither factor will act in isolation.
Organic matter
There are conflicting reports about the influence of organic matter on the survival and transport
of micro-organisms in the subsurface, with different responses being noted for bacteria and
viruses, and for different species and strains within each group. The influence of organic matter
on virus survival has not been firmly defined. In some studies it has been found that
proteinaceous material present in wastewater may have a protective effect on viruses; however,
in other studies no effect has been observed. Whilst similar observations have been made of
bacterial survival in the presence of organic matter, there remains an additional concern that
enteric bacteria, in particular the major pathogens and faecal indicator organisms, may be able to
undergo a certain level of growth in the environment if the conditions are suitable. There is some
support for this hypothesis, a few reports have been published demonstrating regrowth of faecal
indicator bacteria in organically rich tropical surface waters, but the evidence is still insufficient
to confirm that regrowth is a significant issue for most enteric bacteria in groundwater.
Dissolved organic matter has generally been found to decrease virus adsorption by competing
for binding sites on soil particles and the aquifer matrix. The consequence of this observation is
that organic matter will increase the mobility of viruses in the subsurface (Powelson et al. 1991).
However, at relatively low concentrations of organic matter the effect may be reversed, causing
increased virus adsorption and significantly reduced mobility in the subsurface (Robertson and
Edberg, 1997).
Overall, bacteria may respond differently. Binding to surfaces is a characteristic of the growth
cycle of most, if not all species of bacteria. Unlike the passive processes that characterise the
attachment of viruses to surfaces, bacterial attachment involves active processes, including the
synthesis of extracellular appendages specifically required to stabilise the bacteria-surface
interaction. The initiation of this binding is favoured by the formation of a conditioning film of
organic molecules deposited on the solid surface (Bitton and Harvey, 1992; Wimpenny, 1996).
Thus, the presence of organic matter may restrict the dispersal of bacteria in the subsurface but
increase their survival time at the site of attachment.
Table 18 below summarises the major environmental factors and their influence on survival and
migration of pathogens.
46
Table 18 Influence of major factors on the survival and migration of micro-organisms in the
subsurface. From Pedley et al. 2005.
Viruses
Factor
Influence on
survival
Influence on
migration
Bacteria
Influence on
survival
Influence on
migration
Temperature
Persistence is longer Unknown
at low temperatures.
Persistence is longer Unknown
at low temperatures.
Microbial
activity
Varies: some viruses Unknown
are inactivated more
readily in the presence
of certain microorganisms,
the
opposite may also be
true, or there may be
no effect.
The
presence
of Unknown
indigenous
microorganisms appears to
reduce the survival
time
of
enteric
bacteria;
possible
synergism with some
protozoa may extend
survival times.
Moisture
content
Most viruses survive
longer in moist soils
and even longer under
saturated conditions;
unsaturated soil may
inactivate viruses at
the air-water interface.
Most enteric viruses
are stable over pH
range of 3 to 9;
however, survival may
be prolonged by near
neutral pH values.
Virus
migration Most bacteria survive
usually
increases longer in moist soils.
under saturated flow
conditions.
Bacterial
migration
usually
increases
under saturated flow
conditions.
Low pH typically Most enteric bacteria
increases
virus will survive longer at
sorption to soils; high near neutral pH.
pH causes desorption
thereby
facilitating
greater migration.
Low pH encourages
adsorption to soils and
the aquifer matrix; the
tendency of bacteria
to bind to surfaces
may reduce the risk of
desorption at high pH.
Soil properties
Probably related to the Greater migration in Probably related to the
degree
of
virus coarse textured soils; degree of bacterial
sorption.
soils with charged adsorption.
surfaces, such as
clays, adsorb viruses.
Greater migration in
coarse textured soils;
soils with charged
surfaces, such as
clays, adsorb viruses.
Association
with soil
Association with soil
generally
increases
survival,
although
attachment to certain
mineral surface may
cause inactivation.
Viruses
interacting
with the soil particles
are inhibited from
migration through the
soil matrix.
Adsorption onto solid
surfaces
increases
survival times; the
concentration
of
bacteria on surfaces
may be several orders
of magnitude higher
than the concentration
in the aqueous phase.
Bacteria
interacting
with the soil particles
are inhibited from
migration through the
soil matrix.
Bacteria/virus
type
Different virus types
vary
in
their
susceptibility
to
inactivation
by
physical, chemical and
biological factors
Sorption to soils is
related to physicochemical difference in
secondary and tertiary
capsid
surface
structure and amino
acid sequence.
Different species of
bacteria vary in their
susceptibility
to
inactivation
by
physical,
chemical
and biological factors.
Some
species
of
bacteria are more
capable of binding to
surfaces;
variation
may
also
occur
between strains of the
same
bacterial
species.
pH
47
Salt species and
concentration
Certain cations may
prolong survival
depending upon the
type of virus.
Increasing ionic
strength of the
surrounding medium
generally increases
sorption.
Unknown
Increasing ionic
strength of the
surrounding medium
generally increases
sorption.
48
Viruses
Factor
Influence on
survival
Bacteria
Influence on
migration
Influence on
survival
Hydraulic
conditions
Unknown
Virus
migration Unknown
generally increased at
higher hydraulic loads
and flow rates.
Organic matter
Organic matter may
prolong survival by
competitively binding
at air-water interfaces
where inactivation can
occur.
Soluble organic matter
competes with viruses
for adsorption on soil
particles, which may
result in increased
virus migration.
Influence on
migration
Bacterial
migration
generally increased at
higher hydraulic loads
and flow rates.
The
presence
of
organic may act as a
source of nutrients for
bacteria,
promoting
growth and extended
survival.
Organic matter may
condition
solid
surfaces and promote
bacterial adsorption.
Modelling approaches
Principally, two different cases of transport can to be distinguished:
1. transport through a porous media like sandstones or gravels (Figure 4 a-d)
2. transport through fractured media like limestone or granite (Figure 4 e-f)
From a physical point of view, both cases involve the same transport mechanisms, namely
advection, dispersion, exchange with the rock matrix and inactivation. However, transport
through fractured media is rather difficult to model because flow directions are often unknown,
flow velocities can vary largely between dry and wet periods and because interaction between
the rock matrix and the fracture are complex. There are analytical solutions available for
simplified cases like contaminant transport through a fracture including and excluding exchange
with the adjacent rock matrix (Tang et al. 1981). These solutions may be used for lab
experiments but they are simplifying far too much to simulate natural conditions. It is therefore
difficult, not to say impossible, to predict transport and occurrence of micro-organisms in karstic
environments. Principally, these environments are rather vulnerable because transport times can
be very short and interaction with the rock is small following the rule “what goes in comes out”.
Even though we probably don’t know the location where it comes out when we introduce
contamination, or on the other hand, where it was introduced when we detect contamination in a
well.
For porous media, analytical solutions are often available for simplified cases, depending on the
boundary and initial conditions. For more complex environments (e.g. river deposits or layered
sediments) numeric codes are available, such as Modflow or Feflow. From a physical point of
view, three different approaches can be found in literature to describe transport of microorganisms. They will be discussed in the following section:
Advection, dispersion and equilibrium sorption
The transport equation for constant dispersion and advection coefficients, including retardation
and a sink/source term is
R⋅
∂C
∂ 2C
∂C
= D⋅ 2 −v⋅
+Q
∂x
∂t
∂x
D = α L ⋅ v + D0
[2]
[3]
C: tracer concentration
t: time [days]
z: distance [m]
R: retardation coefficient [-]
D: dispersion coefficient [m2/day]
D0: diffusion coefficient in distilled water [m2/day]
49
v: average flow velocity [m/day]
Q: sink or source term (positive for source, negative for sink)
αL: longitudinal dispersivity [m]
An often-used equation to interpret lab or field experiments is assuming linear equilibrium
sorption that is expressed by retardation and degradation or inactivation only. In this case the
sink term becomes
Q = −λ ⋅ R ⋅ C
[4]
λ: degradation or decay constant [day-1]
For this equation, a series of analytical solutions are available for different initial and boundary
conditions (Kreft and Zuber, 1978; Parker and van Genuchten, 1984). The retardation factor is
then used to adjust the peak concentration or the breakthrough to the observations. This
approach can lead to peak arrivals of the micro-organisms sooner than a conservative tracer,
simply due to the inactivation of the contaminant (Figure 8). Such interpretations of results has
led to certain confusions because they imply that the micro-organisms are moving faster than a
conservative tracer, representing the average groundwater velocity. Retardation factors <1 are
known from colloid filtration theory and were first observed by (Small, 1974). He found
retardation factors of minimally 0.9 for large particles and called it hydrodynamic
chromatography. In groundwater modelling, this effect is often called “velocity enhancement”
because particles are moving faster than the average water velocity. Note, that a documented
retardation factor smaller than one therefore does not necessarily mean that velocity
enhancement occurs. More likely, it is an artefact of the modelling approach used.
30
C
25
20
15
10
5
0
0
5
10
15
t [days]
20
Figure 8 1-dimensional analytical solution of tracer concentration at x=6m away from the injection
point, average groundwater velocity of 1m/day and a dispersion of 2m2/day. The full curve indicates
the solution for l=0day-1 and the dotted curve shows the result for l=0.5day-1. The vertical lines
show the time of peak arrival at the observation point.
The second approach assumes equilibrium sorption, where S is the mass of contaminant that is
adsorbed on the aquifer matrix. It can as well be expressed as a sink term in equation 2.
50
Q = −(1 − n ) ⋅ ρ B ⋅
∂S
∂t
[5]
For instantaneous sorption, the partitioning between dissolved and adsorbed contaminant can be
expressed using different sorption isotherms:
Linear, reversible sorption:
S = K ⋅C
[6]
This approach can easily be introduced into equation 5 and then into equation 2, leading to a
retardation constant R of
R = 1 + (1 − n ) ⋅ ρ B ⋅ K
[7]
However, sorption and desorption are often not linear. Therefore, other approaches are used to
simulate sorption more accurately.
Langmuir’s isotherm:
K1 ⋅ C
S=
1+ K2 ⋅C
Freundlich’s isotherm:
S = K1 ⋅ C 1 / K 2
[8]
[9]
Non-equilibrium Sorption
In reality, kinetic adsorption is usually observed, where the sink/source term Q depends on both
dissolved and adsorbed concentrations and time (Schijven et al. 2000). In this case, the problem
is a bit more complicated because the evolution of dissolved pathogens depend on the amount
adsorbed and vice versa.
ρ
∂C
∂ 2C
∂C
− k att ⋅ C − µ l ⋅ C + k det ⋅ B ⋅ S
= D⋅ 2 −v⋅
∂x
n
∂t
∂x
ρ B ∂S
ρB
ρB
⋅
= k att ⋅ C − k det ⋅
⋅ S − µs ⋅
⋅S
n ∂t
n
n
[10]
[11]
with
C: concentration of free micro-organisms
S: concentration of attached micro-organisms
n: porosity [-]
katt: attachment rate coefficient [day-1]
kdet: detachment rate coefficient [day-1]
µ1: inactivation rate for free micro-organisms [day-1]
µS: inactivation rate for attached micro-organisms [day-1]
ρB: dry bulk density [kg/m3]
The above equations are usually sufficient to reproduce lab or field observation. However, they
do not link sorption rates with physical conditions of water and/or the rock matrix. Therefore,
these equations are only of limited use to predict the transport of a specific micro-organism. To
estimate the above model parameters, more sophisticated approaches, usually used to model
colloid transport, can be used.
The existing codes for virus transport can be placed into two categories. As shown in Table 19,
the first group contains computer codes which are readily available to the public and which have
user's manuals. The second group, shown in Table 19 and Table 21, contain computer codes
which were developed for research purposes. Better understanding of virus transport
mechanisms was the main motivation in developing these codes, rather than public
dissemination.
VIRAL T, developed for EPA's Office of Drinking Water, is a modular, semi-analytical and
numerical code that simulates the transport and fate of viruses in ground water. The code
computes viral concentrations in extracted water describing both steady state and transient
51
transport including advection and dispersion in the vertical direction in the unsaturated zone.
Along ground-water flow lines in the saturated zone it handles adsorption and inactivation.
Table 19 Publicly available virus transport codes. From Azadpur-Keeley et al. 2003.
Program Name Year Authors
VIRVLO
v. 1.0
2002
VlRALT
v. 3.0
1994
CANVAS
v. 2.0
1994
VIRTVS
v. 1.O
1991
Description
Remarks
A Monte Carlo-based screening model
Developed
Faulkner et al. for predicting total virus mass
By EPA@
attenuation in the unsaturated zone.
ORD
V.S. EPA-ORD Processes considered: advection,
dispersion sorption inactivation and
A modular semi-analytic and numerical
Park et al.
code for transport and fate of viruses in Developed
@
the unsaturated zones. Processes
For EPA
Hydro-Geologic considered: advection, dispersion,
sorption and inactivation
A modular semi-analytical and
Descendan
Park et al.
numerical code for transport and fate of t
@
viruses in the unsaturated and saturated OfVIRAL
Hydro-Geologic zones. Processes considered: advection,
T
dispersion sorption
A numerical code for transport and fate
Yates et al.
of viruses in the unsaturated zone. The Research@
virus transport is coupled with the flow Oriented
V.S.
Salinity of water and heat through soil.
Code
Laboratory
Processes Considered: advection,
dispersion, sorption, and inactivation
Table 20 Other virus transport codes developed for research purposes. From Azadpur-Keeley et al.
(2003).
Reference
Chu et al. 2001
Sim and
Chrysikopoulos,
2000.
Lindqvist et al.
1994
Tan et al. 1994
Solution
Method
Finite
Difference
Finite
Difference
Method
Finite
Difference
Method&
Analytical
Finite
Difference
Processes Considered
Medium
Advection, dispersion, mass- Unsaturated
transfer,
adsorption,
and 1-D
Advection,
dispersion, Unsaturated
adsorption, and mass-transfer 1-D
Advection, dispersion, and non- Saturated
equilibrium sorption
1-D
Advection, dispersion, and Saturated
sorption
(max.
retention 1-D
Advection, dispersion, and Saturated
clogging/declogging
1-D
Hornberger et al.
1992
Analytical
Tan et al. 1992
QuasiAnalytical
Dispersion and sorption
Harvey and
Garabedian, 1991
Analytical
Advection, dispersion, sorption, Saturated
1-D
and filtration
Unsaturated
1-D
52
Reference
Lindqvist and
Bengtsson,
1991
Solution
Method
Analytical
Finite
Tim and
Element
Mostaghimi, 1991
Method
Taylor and Jaffé,
1990
Finite
Element
Method
Matthess et al. 1988 Analytical
Corapcioglu and
Haridas,
1985
Matthess and
Pekdeger, 1981
Vilker and Burge,
1980
Analytical,
Finite
Element
Method
Not Clear
Analytical
Vilker et al. 1978
Analytical
Campbell et al.
1999
Stochastic
Processes Considered
Medium
Saturated
Advection, dispersion, nonSand
equilibrium sorption, and decay
Column
Advection, dispersion, linear
equilibrium, sorption, and first Unsaturated
order decay.
Soil
Program Name: VIROTRANS
Advection, dispersion, sorption,
growth/decay,
and Saturated
shear/filtration. The change in Column
parameter values due to
Advection, dispersion,
Saturated
sorption, and filtration.
1-D
Advection, dispersion, sorption,
decay/growth,
and Saturated
clogging/declogging. Transport 1-D &
equation is coupled with 2-D
nutrient concentration.
Discussion
on
controlling Saturated
factors
for
bacteria/virus medium
Adsorption mass transfer
Batch &
model.
Column
Saturated
Ion exchange/adsorption
Column
Advection,
dispersion, Saturated
inactivation,
adsorption- medium
Transport through the unsaturated soil
Hydrogeological processes in the unsaturated zone are complex and the behaviour of microorganisms is often difficult to predict. Nevertheless, the unsaturated zone can play an important
role in retarding (and in some cases eliminating) pathogens and so must be considered when
assessing aquifer vulnerability, as already described in section 0. Attenuation of pathogens is
generally most effective in the uppermost soil layers where biological activity is greatest. The
presence of protozoa and other predatory organisms, the rapid changes in soil moisture and
temperature, competition from the established microbial community, and the effect of sunlight at
the surface combine to reduce the level of pathogens within this zone. The effect of individual
environmental factors has been discussed in section 0.
The transport of pathogens from the surface into the subsurface requires the presence of
moisture. Even during relatively dry periods, soil particles retain sufficient moisture over their
surface for pathogens to migrate downwards into the subsurface. Under these conditions the
main driving forces will be sedimentation, diffusion and bacterial motility. Within the thin film
of moisture the organisms are brought into close contact with the surface of the particle, thus
increasing the opportunity for adsorption to the particle surface and further retarding movement.
If soil moisture decreases, the strength of the association between the organism and the particle
surface will increase to a point where the organism is bound irreversibly to the surface. Passive
binding to particle surfaces has been observed with some strains of virus, and it is believed that
53
the strength of the bond can immobilise the virus and contain it at the point of interaction. It is
possible that similar interactions occur with other groups of pathogens, but the processes are less
well defined. Bacteria, for instance, synthesise extracellular substances that can enhance their
attachment to surfaces and promote binding, suggesting that the process involves both passive
and active stages. Whether alone, or in combination with the apparently protective effect of
adsorption onto surfaces, soil moisture influences the persistence of micro-organisms, in
particular viruses. In laboratory experiments soil moisture content of between 10 and 15 % was
shown to be optimal for the survival of several strains of enteric viruses (e.g. Bagdasaryan,
1964; Sagik et al. 1978; Hurst et al. 1980a; 1980b).
By contrast, an increase in the moisture content of the unsaturated zone may increase the
vulnerability of the aquifer to pathogen contamination in two ways: By providing rapid transport
pathways and by mobilising adsorbed organisms. During periods of high recharge, for example
during prolonged heavy rain, the intergranular spaces in the unsaturated zone become
waterlogged and provide a hydraulic pathway for the rapid transport for pathogens. Where these
intergranular spaces expand into fissures the downward migration of pathogens can be extremely
rapid. For example, particles ranging in diameter from 0.1-6.0 µm have been found to move
through 20 m of unsaturated chalk in less than three days by passage through horizontal and
vertical fissures (Lawrence et al. 1996). Moreover, the rapid movement of pathogens through
fissures limits the potential for attenuation by adsorption to surfaces in the soil matrix.
In the interval between recharge events, the chemistry of the water in the unsaturated zone will
change as it equilibrates with the soil matrix. In some soil types, these changes may favour the
adsorption of micro-organisms to surfaces in the soil matrix. A change in the ionic strength or
salt content of the surrounding medium, which can occur during a rainfall event, may be
sufficient to cause desorption of the organism allowing further migration into the soil. This
phenomenon has been observed in laboratory experiments and there is evidence to suggest that it
can occur in the field. Furthermore, some workers have noted that the virus particles that have
desorbed from the soil surface have a reduced capacity to re-adsorb when the environmental
conditions become favourable. The implication of this observation is that virus particles that
have been mobilised in the subsurface are unaffected by one of the principal methods of
attenuation and likely, therefore, to be dispersed over a much wider area than would be
anticipated.
The size variability of micro-organisms (Table 15) can, to an extent, control their mobility in the
subsurface. Soil and rock pore sizes are also variable and the two ranges are known to overlap.
Thus in soils that are composed of fine grain particles, typically clayey-silts, the pore space is
sufficiently small (<4 µm) to physically prevent the passage of bacterial and protozoal pathogens
into the subsurface. Filtration has been identified as the principal mechanism for controlling the
migration of Giardia and Cryptosporidium species (Cryptosporidium oocysts: 4-6 µm; Giardia
cysts: 7-14 µm) through these soil types; indeed, experience has shown that up to 99 per cent of
Cryptosporidium oocysts are retained in the upper layers of the soil. However, the isolation of
Cryptosporidium and Giardia from a small, but significant number of groundwater sources in the
US (Hancock et al. 1998) and the UK indicates that the protective effect of the soil layer is
frequently evaded, probably by migration through preferential pathways, or bypassing, for
example, from sewers that are often located below the soil zone.
In summary, maximising the residence times in the unsaturated zone has been proposed as the
key mechanism for eliminating bacteria and viruses (Lewis et al. 1982) and, in general, this
principle is robust. However, there are exceptions, for example:
• The variability in the nature and thickness of the unsaturated zone overlying aquifers means
that the residence times may not always be adequate to attenuate all pathogens. In
particular, during periods of high recharge, an aquifer may be vulnerable to contamination
by pathogens that are transported rapidly through the waterlogged intergranular spaces in
the unsaturated zone.
• Where the flow is intergranular within the unsaturated zone there is greater potential for
contact with the soil/rock particles and hence greater potential for retention, both sorptive
and filtering. However, if excessive loading takes place the filtering effect may lead to a
blocking of the pores. The resulting reduction in hydraulic conductivity may reduce the
54
•
effectiveness of the unsaturated zone to retard contaminants if the clogging forces recharge
water into vertical fissures where rapid downward movement can occur.
The structure of the unsaturated zone is seldom uniform and fissures may exist permanently
or develop in any environment when the unsaturated zone dries out. The presence of
fissures will always increase the vulnerability of the groundwater to contamination from the
surface, and it should be considered that although the soil conditions may facilitate the
adsorption and attenuation of pathogens, the existence of bypass channels may offset the
protective effect of the soil.
Transport through the saturated zone
From the perspective of groundwater management and the estimation of pathogens at the point
of abstraction (receptor), highly fractured and karstic aquifers represent a particular problem. As
discussed above, groundwater flow through fractured systems may be very rapid, and the
potential for micro-organisms to be attenuated by interaction with the aquifer matrix is much
reduced, although not entirely absent. Consequently, the inactivation rate of the pathogen and
the groundwater flow rate will primarily control dispersal in these aquifer systems. Three
referenced studies will help to illustrate the potential for rapid pathogen transport in highly
fractured aquifers:
• The migration of bacteriophage in a chalk aquifer in the South of England was investigated
by Skilton and Wheeler (1988; 1989). They injected three strains of bacteriophage into
piezometers that intersected the water table and then collected samples at different sites
down-gradient to determine the extent of movement. Very high velocities were observed at
one site due to the fact that the majority of the water flow is through fissures, fractures,
solution openings, and cavities. All three phage types were detected 355 m from the
injection site approximately 5 hours after introduction. It is noteworthy that viable phage
were still being recovered more than 150 days after they were injected into the aquifer.
• Mahler et al. (2000) cite the work of Batsch and colleagues who reported the detection of
injected bacteria 14 km from the injection site, having been transported at a velocity of
about 250 m h-1. Mahler’s own studies (Mahler et al. 2000) in a karstic aquifer located in
the South of France have confirmed the very rapid transport of faecal indicator bacteria in
these systems.
• Lee (1993) investigated the contamination of a water supply well by Giardia spp. and
Cryptosporidium spp. in a karstic environment. The karstic nature of the study area
provided the potential for rapid infiltration of surface waters to the water table and
subsequent transport of the organisms to the well through fractures and fissures. This
connection was confirmed by the study. An analysis of particle size revealed that the full
range of particle sizes found in the surface waters was not present in the well; there was a
cut-off at both high and low ranges. The author concluded that there had been adsorption of
smaller particles and straining of larger ones. The size range of the particles that were
transported through the system included Giardia and Cryptosporidium.
These observations have significant implications for the public health risk associated with water
abstracted from highly fractured and karstic aquifers. Not only can viral, bacterial and protozoan
pathogens be transported rapidly over great distances, but also the groundwater flow pattern
between the source and receptor can be very difficult to predict due to the many interconnected
fractures in the aquifer. It is possible that well designed tracer studies and groundwater flow
models can help to define the potential limits of pathogen dispersion in a highly fractured
aquifer; however, with the current uncertainties surrounding pathogen attenuation in
groundwater it is prudent to assume that where these aquifer types are exposed to sources of
pathogens they are at high risk of contamination over a wide area.
Based on their size and longevity in the environment viruses have the highest potential to be
transported to, and within, groundwater. From the data that is available, the same factors appear
to affect the survival of bacteria; however, some bacteria are able to utilise specific physiological
responses to resist environmental stress that are not available to viruses.
55
Review of public domain information on contamination
level of all types of source water in the European Union
There have been a number of studies undertaken to investigate the occurrence of Campylobacter,
Cryptosporidium and Giardia in particular, in source waters. Fewer studies have been published
on the levels of viruses, and E. coli 0157. In all cases presented below it should be borne in mind
that the sampling and testing methods varied and as such can influence the numbers of
pathogens detected. Methods differ in their sensitivity and selectivity, and in vitro culturing
techniques do not isolate all the organisms present in samples, due to the differences in
metabolic condition of individual cells. This has been discussed in section 0. In addition,
analytical methods may be less precise depending on the type of water source, water purity,
turbidity, season and a number of other factors related to transport and survival of the organisms
of interest which have been discussed in sections 0 and 0.
Cryptosporidium
Cryptosporidial oocysts can be identified in almost all surface waters at some time (Tillett et al.
1998). The many studies identifying Cryptosporidium in surface waters have been reviewed in
all three of the reports of the UK Group of Experts on Cryptosporidium in Water Supplies (Anon
1990; Anon 1995; Anon 1998) and by Rose et al. (2002). These reviews and the selection of
studies described below confirm the ubiquitous presence of Cryptosporidium in surface waters
but also highlights the importance of molecular characterisation to identify the source of the
protozoan. What is obvious from the reports is that there is considerable variation in the numbers
of oocysts in source waters. In the UK for example, sampling of six river sources and one
lake/reservoir between the years 2000 and 2003 revealed that the number of oocysts in the rivers
varied 1-2 log orders throughout the year, and the numbers in lakes/rivers were at least 2 log
orders lower. Spikes in the reservoir/lake water were seen which may reflect catchment activity
(G. Stanfield Pers. Comm).
Due to methodological limitations, uncertainty of accurate measurements exist leading to best
estimates of oocyst numbers and distribution in raw waters, and the factors affecting survival
(Gale, 2001). While method improvement has permitted routine monitoring of water supplies for
Cryptosporidium and Giardia, many studies probably still underestimate parasite contamination
due to their intermittent occurrence. Concentration and frequency distribution depends of
fluctuations in source, weather (e.g. leading to peak flow events) and other parameters. The
regular detection of Cryptosporidium spp. in fresh waters has been reported in the Czech
Republic, in connection with the summer floods of 1997 (Dolej et al. 2000), for example.
Sample surveys have shown that the occurrence of (oo)cysts in surface waters is probably
related to land use influences, agricultural activity including effluent, and sewage discharges.
Therefore contamination of surface waters with Giardia and Cryptosporidium may come from
either human or animal sources.
The USA Information Collection Rule mandated collection of data on water quality including
Cryptosporidium in source water in the USA from 1996 to 1998. Of the 5838 samples assayed,
20% were positive. Badenoch (1995) reports the presence of oocysts in all types of surface
waters with figures ranging from 0.006-2.5 oocysts per litre of water. Lisle and Rose (1995)
found Cryptosporidial oocysts in between 4 and 100% of water samples examined in the USA
and UK, at levels of between 0.1 and 10,000 per 100 litres depending on the impact from sewage
and animals. More recently, Smith and Grimason (2003) reported that in over 3700 surface
water samples from 11 countries, 0-100% of specimens contained between 0 and 252.7 oocysts
per litre.
Ward et al. (2002) characterised Cryptosporidium spp. isolated from various types of surface
waters – rivers, creeks, lakes, sewage plant in- and out-lets and swimming pools around Zurich
(Switzerland) and Munich (Germany). Cryptosporidium oocysts were isolated by continuousflow-centrifugation and immunomagnetic separation (IMS) and the genotypes and species were
characterised using PCR combined with direct sequencing of the amplicon which spans a
56
variable region of the 18S rRNA. Cryptosporidium spp. were detected in 23 of the 68 water
samples investigated. Almost half of these isolates represent species and genotypes known to be
pathogenic to man, namely C. parvum 'bovine genotype' and C. parvum human genotype. Other
genotypes were also detected – C. muris, A and B, C. baileyi as well as three new genotypes.
Extensive monitoring programmes in the Netherlands in the late 1990s showed that the water
quality of the River Rhine and River Meuse is strongly influenced by domestic and industrial
wastewater discharges and agricultural runoff (van Breemen et al. 1998). Medema et al. (1996)
showed that Cryptosporidium and Giardia densities in the River Meuse and River Rhine were of
similar orders of magnitude, with the highest densities in the Belgian part of the River Meuse
(Tailfer; Table 21).
Table 21 Mean Cryptosporidium and Giardia densities in the rivers Rhine and Meuse in 1995 (from
Medema et al. 1996; Values corrected for recovery of detection method).
Location
Number
Cryptosporidium
of samples
(oocysts/l)
River Meuse
Tailfer (Belgium)
4
34
Eijsden
(Belgium/
4
5.3
Netherlands)
Keizersveer
4
4.1
(Netherlands)
River Rhine
Lobith
(Germany/
5
4.5
Netherlands)
Zwolle (Netherlands)
5
4.3
Nieuwegein
5
12
(Netherlands)
Giardia (cysts/l)
94
95
19
22
24
13
A follow up study undertaken between 1997 and 1998 showed that the annual protozoan load in
the Rhine at Lobith was 1.7 x 1015 of Cryptosporidium oocysts and 1.4 x 1015 of Giardia cysts; in
the Meuse near Eijsden this was 2.1 x 1014 of oocysts and 3.6 x 1014 of cysts (Hoogenboezem et
al. 2001). The source of these are not confirmed but it is thought that the contributions made by
treated and untreated municipal wastewater and effluent of calf manure processing are small,
and that import through international rivers could be a significant source. The (oo)cysts present
in the manure of calves, veal calves and commercial egg layers could potentially make a
significant contribution, especially in local surface waters not influenced by the Rhine or Meuse
but it is not known what percentage of this manure reach Dutch surface waters.
In Russia, raw river waters throughout the European Russian region were found to contain both
Giardia and Cryptosporidium with Cryptosporidium oocysts detected in 23/87 (26%) while
26/87 (30%) source surface waters throughout the region contained Giardia cysts (Egorov et al.
2002).
A total of 432 samples were analysed for Cryptosporidium in 25 public water supplies in Ireland
in 2002. The majority of the samples relate to the monitoring carried out in response to the
outbreak of cryptosporidiosis associated with the Mullinger supply, Co. Westmeath, which
occurred in April 2002. The supply is a chlorinated spring-fed lake serving approximately
15,000 persons. A total of 29 cases of cryptosporidiosis were confirmed during the outbreak
(Garvey et al. 2002). Cryptosporidium was detected in concentrations of up to 2.4 oocysts per 10
litres during the outbreak. Cryptosporidium was not detected in any other supply that was
monitored in 2002.
There appears to be strong seasonal variation in the occurrence of Cryptosporidium in surface
waters (see for example, Bodley-Tickle et al. 2002). In an on-farm study in the UK conducted by
Bodley-Tickell et al. (2002), Cryptosporidium was detected in surface waters throughout the
year but with the highest frequency and maximum concentrations during the autumn and winter,
57
coinciding with calving and peaks in wildlife populations, but not with rainfall or slurry
spreading. These authors also studied a small pond at the top of a catchment which was not
under the influence of livestock, in which the only source of oocysts detected could have been
wildlife. Other studies, however, have detected a link between climatic factors and (oo)cysts
concentrations (see Rose, 2002).
LeChevallier et al. (2003) analysed Cryptosporidium occurrence in six watersheds in the USA
using two different methods – methods 1623 and integrated cell culture-PCR. There are two
important points made by this research – one that the two methods used provided comparable
results and the other is that the results suggested that most surface water systems would require,
on average, a 3-log reduction in source water Cryptosporidium levels to meet potable water
standards.
There have been a number of reports of cryptosporidiosis which have implicated groundwater as
a significant source. Hancock et al. (1998) report between 9.5% and 22% of groundwater
samples in the USA positive for Cryptosporidium, although at low concentrations.
In November 1992 and February 1993 in the North West of England 47 cases of
cryptosporidiosis were reported. There was a strong association between cases and residence in
an area supplied from two groundwater sources. Although oocysts were not detected in the water
supply it was found that during heavy rainfall one source was found to drain surface water
directly from a field containing livestock faeces, bypassing natural sandstone filtration. It was
therefore concluded that the groundwater had been contaminated in this way (Bridgman et al.
1995).
The first published case of a Cryptosporidium outbreak caused by filtered borehole water is
reported by Willcocks et al. (1998). In North Thames, UK in the spring of 1997, 345 confirmed
cases of cryptosporidiosis were reported. The detection of oocysts in the water (concentrations
not published), together with the descriptive epidemiology, attack rates, and a case control study,
suggested that the outbreak was associated with drinking water that originated from a deep chalk
borehole.
During June 1996, water supplies of San Pedro Sula, Honduras, were sampled to assess levels of
Cryptosporidium and Giardia. Each sample was concentrated and strained with an indirect
immunofluorescent antibody, and parasites counted through microscopic analysis. In three
surface water supplies, Cryptosporidium oocysts concentrations ranged from 58-260 oocysts per
100 L, and Giardia cysts were present in concentrations ranging from 380-2100 cysts per 100 L.
Groundwater samples had higher concentrations of Cryptosporidium oocysts (26/100 L) than
Giardia cysts (6/100 L). The authors (Solo-Gabriele et al. 1998) suggested that this indicated the
groundwater aquifer was protecting the water supply more effectively from larger Giardia cysts.
According to these authors the concentrations of Cryptosporidium oocysts recorded are in the
typical range for surface water supplies in North America whereas the Giardia concentrations are
elevated.
The two main rivers of Ile-de-France: the Seine and the Marne, are almost exclusively the source
of drinking water supply for five million inhabitants. These rivers are particularly exposed to
pathogens as they flow first through rural areas planted with grain but where cattle grazing is not
significant, and then go into areas of high population density. Rouquet et al. (2000) looked at the
concentrations of Cryptosporidium and Giardia at the drinking water treatment plants’ intakes.
Results showed low levels of Cryptosporidium – generally less than 5 oocysts /l. Giardia
concentrations were more variable. High concentrations of Giardia (>2500 cysts/l) were
observed at the beginning of September in the Marne. The increase was thought to be linked to
storms that occurred at the same time. The study confirmed that Cryptosporidium pollution is
less urban than for Giardia (States et al. 1995) and that summer storm run-off is a possible and
significant source of contamination, especially after a dry period.
Although it is confirmed that Cryptosporidium is widely present in drinking water sources, the
public health significance of oocysts when detected in environmental samples must be
investigated. Not only can the sampling and testing methods heavily influence the numbers of
(oo)cysts detected but Xiao et al. (2001) is amongst the researchers who have shown that a
variety of species and genotypes are detected in surface and wastewaters, some of which are not
known to be infectious for humans. This has implications for both monitoring and for outbreak
58
investigations. Laboratory application of molecular techniques is helping to determine the
sources of the parasites.
Manure management programmes to reduce runoff from agricultural sites, testing and repair or
elimination of leaking septic tanks, and improved cleaning of wastewater that is released into
surface waters will help reduce loadings of all waterborne pathogens including Cryptosporidium
(Fayer, 2004).
Finally, most, but not all, drinking-water outbreaks of Cryptosporidium have been associated
with unfiltered surface water supplies (Hunter et al. 2005). This implies that if filtration is not
100% effective, populations supplied by surface water sources are more likely to be exposed
recurrently to the risk of cryptosporidiosis than those exposed to groundwater (Hunter and
Quigley, 1998).
Giardia
There have been a number of studies which have identified Giardia in source waters in both and
urban and rural setting, confirming their ubiquitous nature.
LeChevallier et al. (1991) studied waters in 14 American states and 1 Canadian province and
found a greater proportion of positive and higher concentration of cysts in urban areas. A
detailed study in British Columbia, Canada found a high sample prevalence of Giardia cysts
(Isaac-Renton et al. 1996). This study was followed up by a longitudinal study of the effects of
watershed management on parasite concentrations (Ong et al. 1996), which showed a similarly
high prevalence, with a distinct seasonal variation, peaking in the winter months. In one of the
watersheds studied, although Giardia cysts were not detected in head water, a mean of 229 cysts
per 100 litres were detected at sample points further down river. Sampling, in relation to
agricultural activity in the catchments, showed a significant increase in the number of cysts
detected from sample points below areas of cattle ranching than above. However, genotypic and
phenotypic variation was observed between cattle, water and human isolates within the
catchment indicated isolated heterogeneity and multiple sources.
Giardia cysts have been detected in >2350 samples of surface water from eight countries at
nearly 5 cysts per litre and in between 21% and 100% of samples examined (Smith and
Grimason, 2003).
Medema et al. (1996) reported Giardia cyst concentrations of 10-100 per litre in samples from
the rivers Rhine and Meuse in the Netherlands.
Roache et al. (1993) report one third of surface waters in remote rural areas of Canada to be
contaminated with Giardia. Wallis et al. (1996) found 21% of raw waters across 72
municipalities in Canada to be contaminated with Giardia, but with wide geographical
(watershed/catchment) variation.
Ongerth et al. (1989) studied three pristine streams in the Pacific Northwest region of the USA
for Giardia cysts using membrane filtration-immunofluorescence assay. Cysts were detected in
43% of samples, with concentrations ranging from 0.1-5.2 cysts per litre. The conclusion from
this study was that Giardia cysts are continuously present, at low concentrations, even in
relatively remote and apparently unpolluted water sources.
Egorov et al. (2002) found 30% of 87 source surface waters (River Sheksna) throughout the
European Russian region contained Giardia cysts at a mean of 2 oocysts per 100 l in raw water
while 7% of finished water contained a mean of 1.6 per 10,000 litres.
Groundwaters have been shown to be vulnerable to contamination by Giardia cysts when they
are under the influence of surface water or other sources of contamination (Moulton-Hancock et
al. 2000).
Campylobacter
As seen with other pathogens, the studies found in the literature reveal different concentrations
of Campylobacter in source waters, depending on location and method of analysis used. In
addition, Campylobacter has been shown to have a seasonal variation, where higher counts are
seen in late autumn and winter, as illustrated by some of the studies described below.
59
Drinking water outbreaks of Campylobacter are usually associated with small, unchlorinated
supplies or drinking from surface waters without treatment, although several reports have
indicated that environmental waters are potential reservoirs and transmitting vehicles for
Campylobacter. Campylobacters have been isolated from bays, rivers, lakes, groundwater and
drinking water. Contaminated drinking water has been the cause of several large outbreaks of
Campylobacter enteritis, particularly in colder countries of the Northern Hemisphere (Mentzing,
1981; Vogt et al. 1982; Taylor et al. 1983). Diergaardt et al. (2004) illustrated this when they
collected various types of water samples (five drinking water, four groundwater, 11 surface
water and four raw sewage) from different parts of South Africa and used Bolton broth to
recover Campylobacter species. Biochemical tests were used to identify Campylobacter isolates.
The study showed lower concentrations of Campylobacter spp isolated from surface waters in
comparison to cooler countries in the Northern Hemisphere.
Feuerpfeil et al. (1997) report more than 10,990 most probable number (MPN)/100mL as the
maximum concentration of Campylobacter in undefined surface water sources.
Bolton et al. (1982) detected between 10 and 36 Campylobacters per 100ml in 7 of 44 river
water samples in the UK using an MPN method. In another study Bolton et al. (1987) looked at
a 15 Km long river system which passed through urban and rural areas of the UK and was
subject to sewage works effluent discharges, over a 12-month period. Using a filtration method,
Campylobacters were found in 43% of 312 samples taken over 12 sites, and using MPN
Campylobacters were found in 21% of samples. The lowest frequency of isolation and the
lowest counts (<10 Campylobacters per 100 ml) were associated with samples from rural sites
and fast-flowing stretches of rivers. The greatest frequency of isolation and highest counts (<10
to 230 Campylobacters per 100 ml) were from sites adjacent to or downstream of sewage works
discharges. Higher counts were seen in late autumn and winter, and fewer isolates seen in spring
and summer. Surface water runoff from adjacent farmland following heavy rainfall was also
found to increase the counts of Campylobacters in the river system.
Knill et al. (1978) collected samples from the Rivers Test and Beaulieu and from two pool sites
near the New Forest in the South of England. Campylobacters were isolated from 74% of the
samples. Fricker and Park (1989) found that 67% of environmental samples from around
Reading, UK, were positive for Campylobacters.
Carter et al. (1987) isolated C. jejuni and C. coli from a number of natural water sources in
central Washington, USA, including ponds, lakes and small mountain streams. At two of the
sites, Campylobacters were recovered throughout the year, although recovery rates were higher
in autumn and winter months.
Stelzer et al. (1989) isolated Campylobacter in 82.1% of river water samples. Levels were
generally below 10 cfu/100ml, but rose to >240 cfu/100ml where waterfowl and faecal
contamination from a poultry farm were present. Stelzer and Jacob (1991) indicated that the
presence of pathogens in surface waters seems to be correlated with water temperature and
incidence of rainfall, whereas Tranter et al. (1996) has correlated their presence with direct input
of contaminants.
Ashbolt et al. (2002) investigated the occurrence of Campylobacter in river waters in six
different catchment areas in Australia and estimated a median concentration of <0.2 to <9.3
MPN/100mL compared with a median of 0.18 and a range of between <0.12 and >11
MPN/100mL detected by Savill et al. (2001).
Jones et al. (1990) sampled surface waters around Lancaster, UK for Campylobacters.
Organisms were absent from an upland reservoir and upper reaches of the River Conder, but
were present in the main rivers entering Morecambe Bay, the lower reaches of the River Conder,
and seawater from the Lune estuary and Morecambe Bay. The surface waters showed
seasonality – higher numbers in winter months and lower numbers in summer. This was thought
to be due to inactivation of the bacteria by solar radiation. As Campylobacter numbers were
lowest when infections in the community were at their peak, it was concluded that contaminated
surface waters were not the reservoir of Campylobacter infections for the community.
Schaffter and Parriaux (2002) looked at the presence of a range of pathogens, including
Campylobacter in surface and groundwaters in mountainous regions in Switzerland.
Campylobacter were detected using classic methods of water filtration on sterile membranes
60
with direct inoculum in enrichment broth. C. jejuni was found to be sporadically present at
almost every site tested. C. coli was less frequent and C. fetus quite rare. As Campylobacter is
unable to grow at temperatures below 30oC, it was concluded that its presence in water must be
related to the presence of contaminant input. The authors looked at potentially contributing
factors and concluded that surface and groundwaters possessing low purification capacities do
episodically contain low levels of potential pathogens. The levels may increase in cases of
flooding. The authors stress that the results obtained for middle and high mountain regions are
probably not transposable to low altitude regions where land use is different and more intensive.
Horman et al. (2004) undertook the first study investigating the presence of Campylobacter spp,
Giardia spp, Cryptosporidium spp., Norovirus and indicator organism in surface waters in Southwestern Finland. A total of 139 water samples were taken from 30 different sites on five
separate sampling occasions in a 2-to-3-week period in consecutive seasons. The sampling sites
included the most important lakes and rivers representing various contamination sources and
catchments areas in south-western and coastal Finland. A total of 41% of samples were positive
for at least one of the enteropathogens analysed. Campylobacter was isolated from 17.3% of
samples, noroviruses were detected in 9.4% of samples; Giardia in 13.7% of samples and
Cryptosporidium in 10.1% of samples. There was some tendency for seasonal variation. There
was no obvious difference between the sites sampled, possibly reflecting the fact that all the sites
were relatively densely populated subject to discharges from human activities, as well as from
agriculture. Although some studies have shown a correlation between faecal indicators and
certain pathogens, this study did not. This may be due to different microbial densities in the
original contamination sources, and therefore the failure to detect pathogens was due to small
sampling volumes.
E. coli 0157
Few studies have been published showing contamination levels of drinking water sources with
enteropathogenic strains of E. coli even though it has become a significant worldwide cause of
illness and transmission has been linked to contaminated water. Jones and Roworth (1996) for
example, report an outbreak of Campylobacter and E. coli 0157 in eight and six people
respectively in Fife, Scotland. Stream water contaminated with treated sewage was blamed for
outbreak which affected the public water supply of a Fife village with a population of about
1100. VTEC 0157 may enter source waters directly via effluent from sewage treatment works,
but this mainly represents those organisms associated with infections in humans, and not those
from the main animal reservoirs.
Schindler (2001) found EHEC during general monitoring of lakes and rivers in Germany. In
1,202 out of 1,347 river samples total E. coli was found in concentrations of more than 200 E.
coli/100mL. These 1202 samples were tested for EHEC, and 31 samples were PCR positive for
EHEC. Similarly, 368 samples out of 4,809 lake samples contained more than 200 E.
coli/100mL, and six out of these 368 samples were PCR-positive for EHEC (Schindler, 2001).
Water treatment processes, especially disinfection, provide effective barriers to transmission of
VTEC 0157, to the extent that outbreaks have not been associated with properly-operated
disinfected public water supplies (Stanfield Pers. Comm). Waterborne VTEC 0157 infections
therefore tend to be associated more with private supplies than with mains water supplies. Where
mains water has been implicated in outbreaks, there is usually circumstantial evidence pointing
to a fault or treatment failure leading to contamination (Chalmers et al. 2000).
Enteroviruses
Numerous studies have shown the presence of enteroviruses in raw and treated water (Keswick
et al. 1984; Gilgen et al. 1995; Reynolds et al. 1997; Chapron et al. 2000; Vivier et al. 2004).
However, because the current methods available for the detection of viruses tend to be long and
complicated there are not many studies that have looked at a large number of samples and most
of the results are therefore from isolated studies. As methods improve and become more rapid
and simple this may change. In addition, the development of a quantitative method for viruses in
water will provide a better estimate of the concentrations of viruses found in source waters.
61
Sampling error, sample size, differences in water flow quantity, differences in techniques and
recovery efficiency between laboratories makes a comparison of numbers difficult.
Human enteric viruses have been shown to persist through sewage treatment processes and be
present in drinking water supplies that meet all the specifications for treatment, disinfection and
counts of indicator organisms (Vivier et al. 2004). Outbreaks of viral gastroenteritis due to
sewage contaminated drinking water have been reported (see for example Hafliger et al. 2000).
Payment et al. (1985) sampled 7 drinking water treatment plants in the US, twice per month for
12 months. Samples were obtained at each level of treatment. It was found that raw water quality
was usually poor, with average virus counts of 3.3 MPNCU/litre and viruses were detected in
7% of finished water samples at an average density of 0.0006 MPNCU/litre.
The particulate phase of river water was also found to contain enteroviruses by Payment et al.
(1988). Sampling of a UK river in 2002 showed that coxsackievirus B2 and B4 were the most
common serotypes identified throughout 2002 but numbers reduced later in the year. During the
summer coxsackievirus B5 numbers increased. Coxsackievirus B1 isolates became increasing
numerous after the middle of the year and Coxsackievirus B3 was present in low numbers
throughout the year (Percival et al. 2004).
Hoogenboezem et al. (2001) report enteroviruses in the Lobith on the Rhine at loads of 5.8 x
1012 and in The Meuse at Eijsden at 1.2 x 1012 per year. Theunissen et al. (1998) sampled
different types of drinking water sources of the Netherlands for enteric viruses and found 0.3-4
L-1 (maximum of 13L-1) in river waters, <0.003-13L-1 in dune filtrate and <0.033-13L-1 in
river filtrate.
De Roda Husman et al. (2004) conducted a study to measure the surface water quality of the
Rhine at two places with respect to human pathogenic viruses. Cultivatable viruses were
detected using membrane filtration and ultrafiltration and then eluted using beef extract. Bovine
green monkey kidney cells were inoculated with part of the concentrate to demonstrate
enteroviruses and reoviruses. Non-cultivatable viruses such as norovirus, were detected using
molecular methods such as PCR. Infectious enteroviruses were found in 90% of samples
ranging from 0.0033 to 0.46 plaque forming particles per litre of water in April and December,
respectively. Norovirus RNA was detected in four out of ten samples taken from the Rhine. In
the Meuse catchment area 41% of samples taken from 1999-2001 were positive for norovirus
RNA. Enteroviruses were found to correlate with turbidity of the water.
Other viruses have been found to be present in Dutch surface waters including hepatitis A and E
(Van der Poel et al. 2002). Hepatitis E was recently detected in over 20% of swine farms in the
Netherlands (Van der Poel et al. 2001) and a cluster of Hepatitis E cases in humans was found in
the north of the Netherlands (Widdowson et al. 2003). The cause was not found, but none of the
infected patients had travelled abroad. Sampling of the drinking water did not show any
evidence of faecal contamination.
Between 1970 and 1979, four laboratories in different areas of the German Democratic Republic
analysed 1,908 surface water samples from 30 sites for the presence of enteric viruses.
Coxsackievirus (particularly types B3 and B4) was isolated every year during the sampling
period (Walter et al. 1982). Treated sewage discharges into the river were not disinfected.
Lucena et al. (1985) collected samples from the Liobregat River and the Besos River and found
an average coxsackievirus type B concentration of 0.107 and 0.60 most probable number
cytopathic units per L, respectively.
All of the rotavirus outbreaks reported have been associated with direct faecal contamination of
a water supply or suboptimal drinking water treatment. Rotaviruses have been detected in
surface waters worldwide with average concentrations ranging from 0.66 to 29 per litre (Gerba
et al. 1996). The highest concentrations have been reported in surface waters receiving untreated
sewage discharges.
Enteroviruses have occasionally been detected in groundwater (Slade, 1981; Abbaszadegan,
1993; 1998; 1999) and in wastewater-recharged groundwater (Vaughn et al. 1978) and there has
been concern that this may affect the quality of the source water for drinking water abstraction.
Powell et al. (2003) showed that an urban aquifer in the UK was contaminated by enteroviruses
during different times of the year.
62
Norovirus
Historically, norovirus have been under-reported because the virus is hard to cultivate. With the
development of specific and sensitive reverse-transcriptase PCR, noroviruses are recognised as
increasingly important causes of gastroenteritis in all age groups (Kirkwood, 2004).
In 2002 an increase of 77-126% in norovirus outbreaks was reported in the Netherlands, England
and Wales, and Germany compared with previous peak seasons in 1995, 1996 and 1999 (Noel et
al. 1999). A new strain of the Genogroup II.4 genotype predominated. Although there is no
evidence that the outbreaks are waterborne, the data illustrates the increase in this virus in
Europe.
Horman et al. (2004) undertook the first study investigating the presence of Campylobacter spp,
Giardia spp, Cryptosporidium spp., Norovirus and indicator organism in surface waters in
Southwestern Finland. A total of 139 water samples were taken from 30 different sites on five
separate sampling occasions in a two-to-three-week period in consecutive seasons. Noncultivatable viruses such as norovirus, were detected using molecular methods such as PCR.
Infectious enteroviruses were found in 90% of samples ranging from 0.0033 to 0.46 plaque
forming particles per litre water in April and December, respectively. Norovirus RNA was
detected in four out of ten samples taken from the Rhine. In the Meuse catchment area 41% of
samples taken from 1999-2001 were positive for norovirus RNA. Enteroviruses were found to
correlate with turbidity of the water.
Hot et al. (2003) sampled four French rivers monthly or semi-monthly for the qualitative RTPCR detection of enteroviruses, Norwalk virus I viruses, Norwalk virus II viruses and other
viruses genomes over a 12-month period. Norwalk-like virus genogroup II was demonstrated in
1.5% of the 68 water samples tested. Because of the genetic diversity of Norwalk-like strains,
the authors concluded that the RT-PCR assays might be unable to detect some strains, therefore
contributing to the very low detection rates of enteric pathogen viruses in the water samples
tested.
Kukkula et al. (1999) identified norovirus in a municipal water system serving a village in
central Finland. The water source is Lake Kermajarvi. The sewage of the village (Heinavesi) is
released into a lake downstream. Karvio village, in the northern part of the municipality, is
located about 6 Km upstream from Heinavesi centre and the water supply comes from private
wells without any municipal sewage system. Norovirus was detected by RT-PCR in the
untreated water, the treated water, and in tap water samples from different parts of the network.
The source of the contamination is unknown. The closest sewage effluent released into the lake
was from the service station with a restaurant in Karvio; but the possibility of a more distant
pollution source was also considered – four months previously a large food borne norovirus
outbreak affected a large number of school children in a city located 70 Km upstream from
Heinavesi. The virus showed an identical amplicon sequence with that detected in Heinavesi.
During the time between the outbreaks the lakes and rivers were covered by ice, which has been
shown to help the virus survive (Dahling and Safferman, 1979).
Norovirus has been linked to a number of groundwater-related outbreaks (Meschke and Sobsey,
2003). Beller et al. (1997) was the first to show that noroviruses came from a contaminated well
in connection with an epidemic. Powell et al. (2003) showed that an urban aquifer in the UK
was contaminated by noroviruses during different times of the year.
63
Table 22 Summary of concentrations of selected pathogens in water bodies.
Pathogen
Type of
water body
Pathogen
concentration
range
Country
Reference
Cryptosporidiu
m
Surface water
0.006-2.5 oocysts
per litre of water
0.1-10,000 per 100
litres
0-252.7 oocysts per
litre
4.1-12 oocysts per
litre
34 oocysts per litre
UK
Badenoch,1995
USA and UK
Lisle and Rose
11 countries
2.4 oocysts per 10
litres
380-2100 cysts per
100 litre
<5 oocysts per litre
Ireland
Smith and
Grimason, 2003
Medema et al.,
1996
Medema et al.,
1996
Garvey et al., 2002
Not specified
Surface water
River water
Spring-fed lake
Surface water
River
Giardia
River
France
Canada
Surface water
River
10-100 per litre
Netherlands
Sreams
USA
Dune filtrate
0.1-5.2 cysts per
litre
2 oocysts per 100
litre in raw water;
1.6 per 10,000 litres
finished water
10,900 MPN per
100ml
10-36 organisms per
100 ml
<10-240 cfu per
100ml
<0.2-<9.3 MPN per
100 ml
<0.12->11 MPN per
100 ml
>200 per 100 ml
0.0006
MPNCU/litre
0.3-4 per l-1 up to 13
l-1
<0.003-13 l-1
River filtrate
<0.033-13
Netherlands
River
0.0033-0.46 plaque
forming units
0.66-29 per litre
0.0033-0.46 plaque
forming units per
litre
Germany
Surface water
River water
River
River
River
E. coli 0157
Enterovirus
(unspecified)
Honduras
229 cysts per 100
litres
5 cysts per litre
Surface waters
Campylobacter
The
Netherlands
Belgium
River and lakes
Drinking water
treatment plant
River
River
Surface waters
8 countries
Solo-Gabriele et
al. 1998
Rouquet et al.
2000
Ong et al. 1996
Smith and
Grimason, 2003
Medema et al.
1996
Ongerth et al. 1989
European
Russian region
Ergov et al. 2002
Germany
Feuerpfiel et al.
1997
Bolton et al. 1982
UK
Stelzer et al. 1898
Australia
Ashbolt et al. 2002
Australia
Savill et al. 2001
Germany
USA
Schindler (2001)
Payment et al.
1985
Theunissen et al.,
1998
Theunissen et al.,
1998
Theunissen et al.,
1998
De Roda Husman
et al., 2004
Gerba et al., 1996
Horman et al.,
2004
Netherlands
Netherlands
Worldwide
Finland
64
Conclusions
The pathogens reviewed in this report have all high health significance. It is clear that source
waters are contaminated to varying degrees with these pathogens. Their presence and persistence
in water is due to a number of factors, such as survival, transport and control of inputs,
depending on the type of water and on aquifer characteristics in the case of groundwater. Rain
seems to be an important vector in the presence of pathogens in water, and temperature and the
presence of cattle and wildlife are also significant factors. In addition, it is clear that there is
strong seasonal variation in the occurrence of these pathogens in surface waters. Source tracking
techniques are revealing that human sources may be more significant than originally thought.
However, until this is further investigated it is not possible to rank the importance of the sources
of contamination at this stage.
The levels of contamination of source waters has been shown to be strongly influenced by the
methods used to identify the pathogens. Although there has been considerable advances in
analytical methods for the detection of micro-organisms, there is no single method that meets all
the requirements needed to accurately assess the risk of infection from waterborne pathogens or
that gives a true picture of the contamination of source waters. There is therefore a need to
further develop and improve detection methods. More studies are necessary to compare
detection of different pathogens in water sources and infected people or recorded cases to
determine risk levels of certain pathogens.
A number of factors have been shown to influence the persistence of pathogens in water
environments. Temperature, UV light, the predation by other pathogens and their capability of
adsorbing to sediment have been shown to particularly influence the persistence of pathogens. In
order to protect water sources the processes and interactions of pathogens in water bodies need
to be fully understood.
The good capabilities of certain pathogens to persist for prolonged times in the subsurface
should lead to a reconsideration of the correct criteria for the sizing of protection zones, taking
into account the purification capacity of different hydrogeological media for example. In
addition, the relatively common presence of pathogenic bacteria and viruses in drinking water
sources question the use of the usual indicators of contamination. It is not possible to exclude
the possibility of higher levels of contamination during extreme hydrodynamic events such as
high rainfall events or floodings.
In the case of surface water intake and vulnerable springs systematic quality monitoring using
classical indicators should be intensified particularly during potentially critical events.
Furthermore, these analyses should be supplemented with direct pathogen screening to
determine peak contamination levels in surface waters or to demonstrate their absence in
groundwaters during vulnerable periods. These controls should be combined with measures to
limit possible sources of contamination on the catchment in vulnerable zones.
In order to develop criteria for selection of water sources to be monitored and to guide the
monitoring to potentially critical events, it is essential to identify the main sources of
contamination and potential causes of peak events in each body of water to be considered. The
following should be reviewed:
• Magnitude of contamination
• Frequency of the discharge (continuous vs run-off related)
• Type of contamination (animal or human) and
• Distance from the water source and travel time during events.
• Transport and survival properties of potential pathogens.
65
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