Environmental Health Perspectives • volume 118 | number 6 | June 2010 871

Research

Recreational Water Quality 

Criteria

Ambient water quality criteria. Viral, 

bacterial, and protozoan pathogens are 

responsible for infectious disease outbreaks 

among recreators at coastal and inland surface waters (Yoder et al. 2008). To protect the 

public from recreational waterborne illness, 

the U.S. Environmental Protection Agency 

(EPA) established ambient water quality criteria (AWQC; U.S. EPA 1986). The epidemiologic studies on which the current AWQC are 

based suggest that at freshwater beaches, the 

rate of acute gastrointestinal illness attributable to swimming should be about 8 cases 

per 1,000 swimmers when monthly geometric mean (GM) density (concentration) 

of Escherichia coli is < 126 colony-forming 

units (CFU)/100 mL), or if the monthly GM 

density of enterococci is < 33 CFU/100 mL 

(EPA 1984). For marine waters, the gastrointestinal illness rate should be about 19 per 

1,000 swimmers when monthly GM enterococci densities do not exceed 35 CFU/100 mL 

(U.S. EPA 1983). Additionally, single-sample 

maximum values were established to aid in 

day-to-day beach management. For freshwater 

beaches, these maxima are 235 CFU/100 mL 

for E. coli or 61 CFU/100 mL for enterococci; 

for marine beaches, the single sample maximum for enterococci is 104 CFU/100 mL. 

Nonetheless, final implementation guidelines for the 1986 AWQC were never issued, 

and prior to the passage of the Beaches 

Environmental Assessment and Coastal 

Health Act (BEACH Act 2000), only 11 states 

applied the criteria to their marine or Great 

Lakes recreational waters [herein referred to 

collectively as coastal waters (CWs)] (U.S. 

EPA 2006).

Indicators and their limitations. Fecal 

indicator bacteria (FIB), such as enterococci 

and E. coli, are measured because they are 

thought to indicate the presence of fecal 

matter, and potentially pathogens, in surface waters. Compared with the pathogens 

that cause recreational waterborne infections, 

FIB are easier to detect and enumerate using 

well-established and inexpensive methods. 

Epidemiologic studies support the use of FIB 

as predictors of illness rates among swimmers 

(Wade et al. 2003).

Culture techniques for FIB require a minimum of 18–24 hr to perform. Thus, beach 

managers evaluate information that is, at best, 

1 day old when deciding to issue swimming 

bans or advisories. By the time FIB results 

become available, the microbial water quality may have changed substantially (Boehm 

2007; Hou et al. 2006; Whitman et al. 2004). 

Illness rates at beaches that are impacted by 

human sewage are related to FIB, which are 

measured by culture (U.S. EPA 1983, 1984) 

or by rapid molecular methods (Wade et al. 

2006, 2008), but at beaches without point 

sources of human fecal pollution, this may 

not be true (Calderon et al. 1991; Colford 

et al. 2007). If the environmental sources and 

persistence of FIB were similar to those of 

pathogens, FIB should be good predictors 

of human illness. However, FIB may persist or even regrow in soil, plants, sand, and 

sediments (Byappanahalli and Fujioka 2004; 

Byappanahalli et al. 2006; Davies et al. 1995; 

Desmarais et al. 2002)—environments that, 

without a suitable host organism, do not support the replication of human viral or protozoan pathogens. Boehm et al. (2009) recently 

reviewed these and other limitations of FIB as 

indicators.

Inland water recreation, risk, and regulation. Inland flowing (riverine) waters are 

surface waters with a net directional current 

and are confined by banks and stream beds. 

Lacustrine waters are freshwater bodies generally confined to a basin (lakes). According to 

the U.S. EPA ATTAINS database (2009), 

the United States has > 3.5 million miles of 

inland flowing waters and more than 41 million acres of inland waters (IWs).

The 1986 AWQC were based almost 

entirely on epidemiologic studies conducted 

at coastal beaches, with a small portion of 

Address correspondence to S. Dorevitch, UIC 

School of Public Health, 2121 W. Taylor, M/C 923, 

Chicago, IL 60612 USA. Telephone: (312) 355-3629. 

Fax: (312) 413-9898. E-mail: sdorevit@uic.edu

We acknowledge all workshop participants for 

their efforts. Many thanks to S. Briggs, Michigan 

Department of Environmental Quality, for her comments. We acknowledge the Water Environment 

Research Foundation (WERF) and the U.S. EPA for 

organizing the expert workshop.

R.L.W.: Contribution 1583 of the U.S. Geological 

Survey Great Lakes Science Center.

This manuscript has been reviewed by the U.S. 

Environmental Protection Agency (EPA) and by the 

U.S. Geological Survey (USGS). The views expressed 

by the authors are not necessarily policy nor the 

views of U.S. EPA or the USGS.

S.D. currently receives research support from the 

WERF and the Metropolitan Water Reclamation 

District of Greater Chicago. J.A.S. is employed by 

Soller Environmental, Berkeley, CA. The remaining 

authors declare they have no actual or potential competing financial interests.

Received 26 October 2009; accepted 25 January 

2010.

Meeting Report: Knowledge and Gaps in Developing Microbial Criteria 

for Inland Recreational Waters

Samuel Dorevitch,1 Nicholas J. Ashbolt,2 Christobel M. Ferguson,3,4 Roger Fujioka,5 Charles D. McGee,6

Jeffrey A. Soller,7 and Richard L. Whitman8

1University of Illinois at Chicago School of Public Health, Chicago, Illinois, USA; 2Office of Research and Development, U.S. Environmental 

Protection Agency, Cincinnati, Ohio, USA; 3Ecowise Environmental, Penrith, New South Wales, Australia; 4Australian National University, 

Canberra, Australian Capitol Territory, Australia; 5University of Hawaii, Water Resources Research Center, Honolulu, Hawaii, USA; 6Orange County Sanitation District, Fountain Valley, California, USA; 7Soller Environmental, Berkeley, California, USA; 8U.S. Geological 

Survey, Lake Michigan Ecological Research Station, Porter, Indiana, USA

The U.S. Environmental Protection Agency (EPA) has committed to issuing in 2012 new or 

revised criteria designed to protect the health of those who use surface waters for recreation. For 

this purpose, the U.S. EPA has been conducting epidemiologic studies to establish relationships 

between microbial measures of water quality and adverse health outcomes among swimmers. New 

methods for testing water quality that would provide same-day results will likely be elements of 

the new criteria. Although the epidemiologic studies upon which the criteria will be based were 

conducted at Great Lakes and marine beaches, the new water quality criteria may be extended to 

inland waters (IWs). Similarities and important differences between coastal waters (CWs) and IWs 

that should be considered when developing criteria for IWs were the focus of an expert workshop. 

Here, we summarize the state of knowledge and research needed to base IWs microbial criteria on 

sound science. Two key differences between CWs and IWs are the sources of indicator bacteria, 

which may modify the relationship between indicator microbes and health risk, and the relationship between indicators and pathogens, which also may vary within IWs. Monitoring using rapid 

molecular methods will require the standardization and simplification of analytical methods, as 

well as greater clarity about their interpretation. Research needs for the short term and longer term 

are described.

Key words: Clean Water Act, environmental epidemiology, environmental microbiology, indicator bacteria, inland waters, predictive modeling, waterborne pathogens, water recreation. Environ 

Health Perspect 118:871–876 (2010). doi:10.1289/ehp.0901627 [Online 25 January 2010]
Dorevitch et al.

872 volume 118 | number 6 | June 2010 • Environmental Health Perspectives

data coming from Keystone Lake, Oklahoma 

(U.S. EPA 1983, 1984). None of the supporting data were collected at rivers. Additionally, 

European IW recreation research is limited by 

the use of unique settings (Fewtrell et al. 1992; 

Lee et al. 1997), limited groups of participants 

(van Asperen et al. 1998), limited water quality measures (Ferley et al. 1989), and abbreviated reporting of illness rates (Fewtrell et al. 

1994). European trials of randomized exposure to IWs have been conducted as part of 

the Epibathe study (European Commission 

2010; Wiedenmann et al. 2006). However, 

the exposure, three head immersions over a 

10-min period, is different from actual swimming; thus, extrapolating findings to other 

contexts is difficult.

The BEACH Act (2000) mandated that 

the 1986 criteria be applied to all U.S. marine 

and Great Lakes CWs, but the mandate was 

not extended to IWs. This Act also required 

that the U.S. EPA conduct epidemiologic 

studies at beaches to develop information for 

issuing new or revised criteria. As a result, the 

National Epidemiological and Environmental 

Assessment of Recreational Water (NEEAR) 

study (Wade 2006, 2008) has been conducted at marine and Great Lakes beaches. 

To address the BEACH Act requirement 

that the U.S. EPA develop more timely indicators of water quality, the NEEAR study 

used quantitative polymerase chain reaction 

(qPCR) tests for FIB that could produce 

same-day results.

In 2006, the U.S. EPA was sued by the 

Natural Resources Defense Council (NRDC) 

and others for failing to meet BEACH Act 

research and regulatory deadlines for recreational waters. In August 2008, the U.S. EPA 

entered into a consent decree (NRDC/EPA

2008), which mandated that new or revised 

criteria be issued by 15 October 2012. The 

consent decree required that the U.S. EPA 

“[e]valuate the applicability of NEEAR Great 

Lakes data to inland water.” Thus, a policy 

imperative exists to consider establishing IW 

AWQC, but the epidemiologic knowledge 

base for criteria development is very limited. 

Extending AWQC derived from studies conducted at coastal sites to IWs involves major 

assumptions: a) similar densities of FIB reflect 

a similar health risk in inland and coastal settings, presumably because they reflect a similar risk of exposure to pathogens of similar 

infectivity and virulence; b) hydrogeochemical differences among inland lakes, rivers, and 

CWs would be assumed to have nondifferential impacts on the transport and fate of 

indicators and pathogens; and c) the criteria 

derived from the studies conducted at sewageimpacted coastal beaches would be assumed 

to protect against illness in inland settings, 

where the dominant pollutant sources may be 

wildlife and/or agricultural animals.

To assess these assumptions, and more 

broadly, the state of the science that could 

support the application of coastal-derived criteria to IWs, a 3-day workshop of 31 national 

and international experts was convened in 

February 2009 by the Water Environment 

Research Foundation (WERF) with support 

from the U.S. EPA. Workshop participants 

identified critical knowledge gaps and outlined 

research needs that could be met by December 

2010 (the deadline for completing research 

that will be reviewed in the development 

of the 2012 criteria) or between 2010 and 

2015 (for consideration in future AWQC). 

A detailed report of workshop proceedings is 

available online (Water Environment Research 

Foundation 2009). In this meeting report the 

chairs of the working groups and the editor 

of the WERF report have summarized workshop highlights and added updates based on 

subsequent discussions.

CWs and IWs: Superficial 

Similarities

General principles of hydrology, microbiology, 

and public health should apply across all water 

recreation settings. Surface waters, regardless 

of matrix and geologic setting, are governed by 

the same ecologic, hydrologic, and geographic 

principles. Likewise, ingesting a specific quantity of a given viable pathogen in any surface 

water should produce similar health risks. 

Consistent with this notion, health risks associated with water recreation have been identified 

at Great Lakes (U.S. EPA 1984; Wade et al. 

2006, 2008) and U.S. marine settings (Colford 

et al. 2007; U.S. EPA 1983) and with IWs 

in Europe (Ferley et al. 1989; Wiedenmann 

et al. 2006). If monitoring of FIB at inland 

recreational waters was mandated, testing that 

is currently conducted for other regulatory 

purposes could, with sufficient resources, be 

expanded. Thus, at first glance, it seems that 

recreational criteria derived from CW studies 

could be applied to and implemented for IW. 

There are, however, important differences to 

consider between IWs and CWs.

Important Differences between 

CWs and IWs

Several critical differences exist between CWs 

and IWs, which can be understood primarily as a function of the scale of water body of 

interest. Scale here refers to the volume, surface area, related landscape, and the flow (for 

flowing waters). Scale can influence watershed interactions, runoff, dilution, currents, 

wave height, turbulence, resuspension, and 

source complexity.

Differences in pathogen source. The ultimate determinants of health risks are thought 

to be dependent on the dose and virulence of 

pathogens ingested by recreators, not whether 

the waters are categorized as coastal or inland. 

IWs are generally dominated by more rural 

areas and agricultural land use and are thus 

more likely to be affected by wildlife and 

livestock than are coastal watersheds. Warmblooded animals have the potential to carry 

a variety of human pathogenic bacteria and 

protozoa and may pose human health risks 

(Dorner et al. 2004). For example, Leptospira

in the urine of infected wildlife or livestock 

can contaminate surface waters and infect 

humans via skin breaks or by ingestion. U.S.

outbreaks of leptospirosis have occurred in 

the context of IW limited- and full-contact 

recreation (Jackson et al. 1993; Morgan et al. 

2002; Yoder et al. 2008). A recent U.S. EPA 

review identified several recreational outbreaks 

tentatively linked to wildlife and livestock 

sources, although definitive confirmations of 

animal sources have been lacking (U.S. EPA 

2009). With limited dilution in inland settings, bathers themselves can become sources 

of fecal pathogens. Sporadic mild illness 

(Calderon et al. 1991) and, more conclusively, numerous outbreaks of severe disease 

including E. coli 0157:H7 (Bruce et al. 2003; 

Keene et al. 1994; Yoder et al. 2008) have 

been linked to other bathers at IWs.

The importance of sediment in IWs. Small 

lakes and streams are closely associated with 

watershed factors such as soils, runoff characteristics, shoreline processes, and meteorological events. Turbulent flow in IWs may lead 

to resuspension of sediment-associated FIB. 

Numerous studies have found that E. coli

and enterococci can persist and potentially 

regrow in sediments and soils (Byappanahalli 

and Fujioka 2004; Byappanahalli et al. 2006; 

Davies et al. 1995; Desmarais et al. 2002). 

Regrowth of FIB is suspected to occur at the 

water/sediment boundary (Wheeler Alm et al. 

2003; Yamahara et al. 2009). The effect of 

sediment resuspension on FIB in IWs could 

be amplified at the boundary layer, which, 

compared with CW settings, is larger in relation to the volume of surface water. An area 

of consensus among workshop participants is 

that soil and sediments, which are thought to 

contain proportionally fewer pathogens than 

fecal sources, should make larger contributions to indicator densities in surface water 

samples in IWs than in CW.

Differences in hydrogeology that change 

indicator densities. The scale of each IW is 

determined by climatic conditions, geology, 

and ecology for that watershed. The sites for 

land-based contamination of IW are many,

because waterborne pollutants enter IW from 

multiple sites as it flows downstream. Because 

the volumes of water in many IW sites are 

relatively small and land-based pollutants 

are often close to these sites, the dilution of 

pollutants is more limited in streams than 

in coastal settings (Olyphant et al. 2003). 

Parameters such as flow and turbulence vary 
Inland recreational waters—state of the science

Environmental Health Perspectives • volume 118 | number 6 | June 2010 873

substantially within the category of IWs and 

could account for more variability in FIB levels in IWs than in CWs.

The decoupling of indicator and pathogen 

densities in IWs. Human pathogenic viruses 

and protozoan parasites reproduce in the cells 

of their hosts, whereas as noted above, sediments can provide favorable conditions for 

the persistence and re-growth of FIB. A concern among workshop participants was that 

a decoupling (meaning a significantly weaker 

association) of indicators and pathogens may 

occur in IWs. The basis for this concern is the 

combination of the known growth of FIB in 

sediments and the influence that sediment 

is thought to have on IW FIB. This decoupling is likely to result in different indicatorpathogen relationships at coastal and inland 

systems, as well as differences across IWs 

with varying hydrologic characteristics. As a 

result of this decoupling, FIB may overestimate pathogen densities and expected illness 

rates among IW recreators. In part because 

of this putative decoupling, the application

of coastal-derived FIB criteria to inland settings should result in rates of sporadic illness 

(although not necessarily outbreaks of severe 

illness) that are at least as protective in IW as 

they are in CW.

Challenges at Both 

CWs and IWs

Rapid testing methods. Recent epidemiologic 

studies have used quantitative real time polymerase chain reaction (qPCR) measurements 

of enterococci (Colford et al. 2007; Wade 

et al. 2006, 2008), and this method may be 

endorsed in the new 2012 AWQC. Currently, 

qPCR and other rapid methods of measuring FIB are in various stages of development 

(Griffith et al. 2009; Noble and Weisberg 

2005). If the necessary monitoring protocols 

and communications systems were in place, 

these methods could provide same-day water 

quality information in IWs.

One broad concern with the qPCR 

approach is method performance, meaning the precision, accuracy, sensitivity, and 

interlaboratory variability. A second concern 

is the interpretation of qPCR output to estimate health risks. In epidemiologic studies, 

enterococci qPCR results are being directly 

calibrated against rates of illness. That calibration is being performed at coastal beaches 

impacted by treated human sewage. Humic 

and fulvic acids found in sediment (CWs and 

IWs) can inhibit PCR analyses (Rutjes et al. 

2006; Tsai and Olson 1992). Because of the 

importance of sediment in determining surface water quality in IWs, sediment may result 

in more inhibition than occurs in CW. The 

qPCR methods detect viable, nonviable, and 

cell-free DNA of FIB, whereas conventional 

methods detect culturable microbes only. The 

distribution of these components of the qPCR 

signal (Nocker et al. 2007) could be different 

in IW compared with CW, particularly at 

sites impacted by treated wastewater. IW pollutant sources and sediments may result in 

qPCR-illness rate relationships that are different from those described (Wade et al. 2006, 

2008) in CW.

Predictive modeling: opportunities and 

uncertainties in CWs and IWs. Modeling 

approaches offer alternatives to epidemiologic 

studies or extensive microbial monitoring. 

Simple regression modeling of FIB densities 

use real-time information such as meteorological and physical parameters, such as turbidity, to produce a timely and, in the long-run,

a lower-cost alternative to microbiological 

monitoring (Boehm et al. 2007; Frick et al. 

2008; Olyphant and Whitman 2004). 

Regression models are used to issue beach 

notification at three Great Lakes locations 

and on the Schuylkill River (Philly RiverCast 

2009). Mechanistic models, which make use 

of microbial loading, dilution, decay, transport, and other parameters to predict location-specific densities of FIB (Boehm et al. 

2005; Steets and Holden 2003), are probably 

best used for evaluating management practices 

in watersheds. These approaches model FIB 

levels; it is unknown whether they predict 

pathogen presence or illness rates.

The above approaches predict water quality, whereas quantitative microbial risk assessment (QMRA) is used to predict health risks 

in populations (Haas et al. 1999). The inputs 

to QMRA models include readily obtainable 

demographic and water quality data. Health 

risks are predicted using estimates of pathogen 

densities, water exposure, and dose response 

(number of units of pathogens ingested as a 

predictor of illness probability). Conversely, a 

desired water quality target can be modeled to 

meet a health risk target. QMRA allows evaluations of relative risks across a range of sitespecific contamination scenarios. Few studies 

have directly compared QMRA projections 

with epidemiologic observations (Ashbolt 

et al. 1997), although the NEEAR study site 

in Boqueron, Puerto Rico (http://www.epa.

gov/NHEERL/neear/) will prospectively collect information needed to compare modeled 

(QMRA) and observed risk. Predictive models of water quality and QMRA can be only as 

accurate as their inputs; some input data, such 

as dose response, remain limited. Sensitivity 

analyses can evaluate sources of uncertainty 

in model estimates to allow prioritization of 

additional data collection (such as spatiotemporal variability in indicators and pathogens).

Policy and implementation challenges. To 

protect the public from waterborne illness at 

IWs, criteria will have to be established based 

on a targeted level of risk. In the absence of 

stakeholder input, such a targeted risk could 

not yet be considered an acceptable risk. The 

unadjusted rate of illness attributable to swimming at Great Lakes point source–impacted 

bathing beaches appears to be about 20–25 

cases of gastrointestinal illness per 1,000 swimmers (Wade et al. 2008). In addition to the 

rate of illness, the severity of illness attributable 

to water recreation is also an important consideration in characterizing risk. The U.S. epidemiologic studies (set in CW) have described 

rates of gastrointestinal illness, generally 

thought to be mild and self-limited. By contrast, reported disease outbreaks in IWs have 

included rare but potentially life-threatening 

infections, likely due to limited dilution and 

proximity to fecal sources (including other 

bathers). Once elements of acceptable risk,

rate, and severity have been defined, specific 

values of FIB can be evaluated with a goal of 

keeping risk below those levels.

AWQC developed to protect recreational 

water users are applied to other Clean Water 

Act (CWA) programs, such as the listing of 

impaired waters and discharge limits (under 

Section 303d of the CWA). A better understanding of wastewater treatment effects on 

the components of the qPCR signal (viable 

bacteria, nonviable bacteria, and cell-free 

DNA) will help define the value of qPCR 

monitoring for these other CWA purposes. 

Translation factors or a refinement of qPCR 

assays to identify viable cells (Nocker et al. 

2007) may be needed to determine, for example, if treated wastewater effluent met discharge standards. The costs of implementing 

qPCR requirements of new AWQC for both 

CWs and IWs will be significant and could 

result in the allocation of local funds away 

from other water quality programs. Sameday measures of FIB have limited benefits 

beyond beach notification. Continued use of 

culture-based methods in those contexts is 

reasonable. A potential advantage of qPCR is 

the ability to differentiate human from nonhuman sources of FIB. Work published after 

the workshop evaluated numerous promising approaches for both rapidly evaluating 

microbe density and differentiating human 

from nonhuman sources (Griffith et al. 

2009). However, limitations of specific qPCR 

approaches for differentiating sources has been 

described (Stapleton et al. 2009).

New monitoring requirements for recreational waters may encompass the vast number of inland lakes and miles of rivers within 

individual states. The application of predictive models, sanitary surveys, and QMRA to 

develop site-specific standards, particularly 

where the dominant source of fecal pollution 

is nonhuman, is of interest to the regulated 

community. It remains an open question 

whether these alternatives to epidemiologic 

studies would provide accurate and sufficiently 

precise projections of FIB and health risk.
Dorevitch et al.

874 volume 118 | number 6 | June 2010 • Environmental Health Perspectives

Critical Research Questions

Four groups of critical questions should be 

answered regarding measures or models of 

water quality as a means of assessing recreational 

waterborne illness risk in IW (Appendix 1):

1. Microbial indicators as predictors 

of pathogen exposure and health risks. To 

advance our ability to model health risk, we 

must characterize the transport, survival, 

fate, and re-growth (for bacteria) of indicators and pathogens in flowing and nonflowing 

IW. Persistence of pathogenic bacteria needs 

to be better understood, as Campylobacter, 

Salmonella, Shigella, and shigatoxin-producing 

E. coli have been found on algae growing in 

surface waters (Ishii et al. 2006). Similarly 

E. coli O157:H7 can persist in sediments 

(Bruce et al. 2003). Although the dynamics of FIB have been studied in some coastal 

contexts (Boehm et al. 2002; Whitman and 

Nevers 2004), FIB spatiotemporal variability 

and its determinants need to be characterized 

in hydrologically diverse IW.

2. Fecal pollution sources as predictors 

of pathogen exposure and health risk. It is 

important to determine if the source of fecal 

pollution modifies the indicator-health association. The assumption that human fecal pollution presents the greatest health risk needs 

further evaluation. Prior discussions, including 

those at the workshop, focused on rates of illness seen among swimmers at beaches, generally mild and self-limited. Future work should 

also consider illness severity, which may be 

substantial if agricultural animals, wildlife, 

or other bathers are sources of pathogens at 

waters where dilution is limited. The health 

risks to recreators at IW impacted by confined animal feeding operations have not been 

the subject of epidemiologic studies but are a 

potential concern. The possibility of zoonotic 

waterborne viral infections should be investigated, as recent evidence supports possible 

zoonotic origins of some human rotaviruses 

(Banyai et al. 2009; Matthijnssens et al. 2009).

3. Molecular methods for water quality testing. Sediments, which likely contain 

qPCR inhibitors, FIB (viable and nonviable),

and cell-free FIB DNA present a challenge 

to IW monitoring. It is important to know 

how insolation, water chemistry, wastewater 

treatment, hydrologic parameters, season, and 

water matrix differentially affect these components of the qPCR signal. Optimizing the 

primers and probes, particularly those that 

differentiate human from other sources, and

establishing procedures for minimizing naturally occurring PCR inhibitors are priorities.

4. Other approaches to predicting IW 

recreation health risks. Procedures for optimizing and validating predictive models, 

sanitary surveys, and QMRA approaches are 

important for both coastal and IW. The U.S. 

EPA Great Lakes survey tool (http://www.

epa.gov/waterscience/beaches/sanitarysurvey/) 

should be modifiable for IW. For QMRA and 

mechanistic models, inputs such as indicator 

and pathogen concentrations, pathogen loading estimates, and health risks posed by various animal species are needed. An uncertainty 

in estimating the health risk associated with 

livestock is the high variability of protozoa 

excretion rates (Ferguson et al. 2009).

An IW Research Agenda

To provide answers to the questions identified 

in Appendix 1, short-term research (within 

2 years) and longer-term projects (2–5 years) 

are needed. These questions can be addressed 

through targeted literature reviews, computer modeling, field sampling for environmental microbes, laboratory research on 

analytic methods, and human health studies. Interdisciplinary studies would compile 

sediment, soil, hydrology, microbiology, and 

health data, all of which could be used to 

identify predictors of health risk. Other than 

epidemiologic and QMRA research, these 

studies would not directly inform the establishment of AWQC but could advance water 

quality modeling and our understanding of 

sources of risk and uncertainty. Research 

agenda elements are listed in Table 1. Limited 

explanations of several items follow.

Short-term research. Molecular tests for 

FIB. Rapid tests that are strongly correlated 

with pathogen densities are needed to support the development of improved predictive 

models of health risk. Pretreatment with propidium monoazide (PMA) may allow the differentiation of DNA from intact viable cells, as 

opposed to extracellular DNA or DNA in cells 

without a functioning membrane (Nocker et al. 

2007). This PMA-qPCR approach has been 

used to demonstrate a faster decay of the qPCR 

signal of Bacteroidales compared with conventional qPCR (Bae and Wuertz 2009). PMA 

pretreatment or other approaches for quantifying elements of the overall qPCR signal should 

be evaluated for their ability to improve qPCR 

predictions of pathogen presence.

Predictive modeling of health risk.

Comparison of retrospective QMRA analyses 

with previously conducted epidemiologic studies could lead to revisions in QMRA model 

assumptions and inputs to bring projected 

levels of risk in line with risk levels observed 

Table 1. Research to address critical IW criteria questions.

Type of research required 

Critical questions

Library computer 

simulation Laboratory Field Study overview

Determinants of indicatorpathogen relationships in IWX X X L: Advance mechanistic modeling of FIB and pathogens, supported by sampling of water, sediment 

and soil in diverse IW. Repeated sampling to characterize the fate, transport, persistence, and 

re-growth

Sources of indicators, 

pathogens, and health risk

X S: Meta-analysis of epidemiologic studies to evaluate fecal pollutant source as a modifier of the 

indicator–health risk relationship

X X S: Optimize and anchor QMRA models based on prior epidemiologic study results

X X S: Develop a sanitary survey tool for use in future IW epidemiologic and QMRA studies

X X X S: Field sampling of feces from agricultural animals and wildlife to determine human pathogenic 

potential and dynamics

X X L: Epidemiologic studies conducted at diverse IW sites, each with a different dominant source of 

fecal pollution (agricultural animals, wildlife, urban runoff, wastewater)

Molecular methods in IW: 

interpretation, standardization

X S: Develop a database of relationships between rapid molecular-based and culture-based 

measures of indicator microbes described in the literature

X X S: Characterize the persistence of specific molecular targets (human vs. other) in a variety of 

environmental settings and wastewater, looking at viable and nonintact cells and cell-free DNA

X X S: Optimize, simplify, and standardize qPCR methods (particularly for source-specific markers) and 

other rapid methods; use these in epidemiologic studies

X X L: Develop rapid methods for concentrating, identifying, and quantifying pathogens in recreational 

waters

Modeling health risk and 

real-time water quality

X X QMRA validation studies, S (retrospective) and L (prospective)

X X S: Measure real-time physicochemical, hydrologic, meteorologic parameters, with microbes

Abbreviations: S, short term (< 2 years); L, longer term (2–5 years).
Inland recreational waters—state of the science

Environmental Health Perspectives • volume 118 | number 6 | June 2010 875

in epidemiologic studies. This anchoring of 

the QMRA predictions will enhance their scientific credibility for predicting recreational 

waterborne illness.

Longer-term research. Modeling water 

quality in real time. Long-term research 

is required to characterize the transport, 

fate, and persistence of microbes and their 

molecular targets in sediments and soils. 

Understanding how these variables change 

as a function of solar radiation, rainfall, and 

biotic and hydrologic factors will advance 

our ability to develop mechanistic models 

for watershed management and set pollutant 

discharge limits.

Pollutant source as a determinant of 

health risk in IW. New epidemiologic studies would fill the data gap that currently precludes directly comparing the FIB-health risk 

relationship in coastal and IW. Ideally, several 

inland sites should be selected, each with a 

different type of dominant source of fecal pollution. Description of both illness rate and 

illness severity will be important to more completely characterize risk. Health data should 

be collected in conjunction with data on indicators and pathogens in surface waters and 

sediments using conventional and emerging 

microbial detection methods. QMRA should 

be performed in tandem with the epidemiologic studies.

Conclusions

We endorse the development of science-based 

criteria to protect the health of those who 

use marine, Great Lakes, and riverine and 

lacustrine recreational waters. We think that 

the distinction of IW versus CW is of less 

importance than more fundamental variables such as the scale of the body of water, 

the source of the pollutant, and the effects 

of sediment, which translate into differences 

in the densities, transport, and fate of indicators and pathogens. Differences in these 

variables between IW and CW may translate 

into weaker indicator–pathogen and indicator–health risk relationships for IW compared with CW. It remains an open question 

whether sediment in IW changes the relationship between enterococci qPCR measures 

and health risk, which has been described at 

coastal beaches impacted by human fecal pollution. A challenge in addressing health risks is 

the imprecision in defining of risk, as frequent 

mild illness (seen in coastal epidemiologic 

studies) may be of less public health concern 

than infrequent severe illness (described in 

outbreaks of disease in IW). We suspect that 

the application of coastal-derived criteria 

should result in rates of sporadic mild illness 

that are no higher (and possibly lower) in IW 

than CW. We are concerned about outbreaks 

of severe disease caused by fecal matter from 

other bathers, wildlife, and livestock. In IW 

with limited dilution capacity and close proximity to sources, outbreaks of severe disease 

may be difficult to prevent by the application of coastal-derived criteria (this was not 

a conclusion of the workshop and represents 

the authors’ views). As critical research questions are answered, a basis will be established 

for developing criteria that would afford a 

similar level of protection in IW as in CW, at 

least for mild sporadic illness. The implementation of microbial monitoring of IW will 

be a challenge to local government agencies. 

Should rapid FIB monitoring, QMRA, sanitary survey data, or real-time modeling prove 

to be predictive of health risks in IW, these 

approaches could be used to protect the public from recreational waterborne illness.

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