chapter 8 data interpretation: stormwater effects...

CHAPTER 8

Data Interpretation

“If you get all the facts, your judgment can be right; if you don’t get all the facts, it can’t be right.”

Bernard M. Baruch

CONTENTS

Is There a Problem?.......................................................................................................................609 Evaluating Biological Stream Impairments Using the Weight-of-Evidence Approach................611

The Process...........................................................................................................................611 Benchmarks ..........................................................................................................................612 Ranking and Confirmatory Studies ......................................................................................615 Comments Pertaining to Habitat Problems and Increases in Stream Flow ........................617

Evaluating Human Health Impairments Using a Risk Assessment Approach .............................619 Deterministic Approach........................................................................................................619 Probabilistic Approach .........................................................................................................619 Example Risk Assessment for Human Exposure to Stormwater Pathogens ......................620

Identifying and Prioritizing Critical Stormwater Sources.............................................................626 Sources of Urban Stormwater Contaminants ......................................................................626 Case Study: Wisconsin Nonpoint Source Program in Urban Areas ...................................628 Use of SLAMM to Identify Pollutant Sources and to Evaluate Control Programs ...........629

Summary ........................................................................................................................................636 References ......................................................................................................................................637

IS THERE A PROBLEM?

Unit 1 (Chapters 1 through 3) described problems associated with stormwater runoff. Unit 2 (Chapters 4 though 8) described the development of appropriate experimental designs that included selecting the components of the assessment process and determining an appropriate level of effort, plus specific sampling and monitoring activities to assess receiving water impacts. Unit 3 (the appendices) includes additional guidance on conducting specific field activities. There are numerous case study examples throughout these chapters showing how the recommended approach has functioned during previous successful projects. In this concluding chapter, these important issues are highlighted for the data interpretation process. Now that an assessment has been conducted, how does one determine whether or not the receiving waters are impaired and, if so, what is the source, or sources, of the impairment?

609

610 STORMWATER EFFECTS HANDBOOK

As indicated in Chapters 2 and 3, there is a variety of receiving water problems that may be associated with stormwater. The specific problems in any area are dependent on many site conditions and objectives. There are many documented cases, previously described, where stormwater has caused detrimental impairments on receiving water uses and goals. Probably the most common problem is associated with stormwater conveyance (flood prevention) caused by increased amounts of pavement in the drainage area. The increased flows, however, are also responsible for many habitat problems related to the increased stream power and associated unstable stream environment. Other common receiving water problems in urban waters are associated with noncontact recreation (linear parks, aesthetics, boating, etc.). The seemingly simple task of preventing floatable debris from being discharged can be very difficult to accomplish. Much of this book has addressed environmental health issues associated with biological uses (warm-water fishery, biological integrity, etc.). In addition to the habitat destruction problems associated with increased flows and increased stream power, contaminated sediment may be a significant causative agent affecting biological uses. Poor water quality obviously can also significantly affect most of the above uses, in addition to interfering with water contact recreation (swimming) and water supply uses. It is unlikely that these human health uses would be appropriate in any waterway located in a heavily urbanized watershed.

The study design is dependent on the expected problems likely to be encountered (see also Chapter 4). Without having that information at the beginning of a study, the initial list of parameters to be monitored has to be based on best judgment. The parameters to be monitored can be grouped into general categories, depending on expected beneficial use impairments, as follows:

• Flooding and drainage: debris and obstructions affecting conveyance are parameters of concern. • Biological life/integrity: habitat destruction, high/low flows, taxonomic composition of existing

aquatic life, inappropriate discharges, polluted sediment (texture, SOD, and toxicants), and wet weather quality (toxicity, bioaccumulation, toxicants, nutrients, DO) are key parameters.

• Noncontact recreation: odors, trash, high/low flows, aesthetics, and public access are the key parameters.

• Swimming and other contact recreation: pathogens, and above listed noncontact parameters, are key. • Water supply: water quality standards (especially pathogens and toxicants) are key parameters. • Shellfish harvesting and other consumptive fishing: pathogens, toxicants, and those listed under

biological life/integrity, are key parameters.

Obviously, there are definite problems in receiving waters that will dictate many components of the sampling program and measures against which the data are to be compared. These problems may be minor if the watershed is relatively undeveloped, but they can be extreme for fully developed urban or agricultural areas. In addition, local use objectives also dramatically affect the definition of a “problem.” In all cases, however, basic receiving water objectives should include safe drainage, noncontact recreation (acceptable aesthetics), and basic biological life objectives. It is unlikely that contact recreation or biological integrity, with the stream being able to support a full mixture of native organisms, would be reasonable receiving goals in a fully developed urban or agricultural watershed.

The information and guidance provided in this book should enable a researcher to investigate local conditions to identify local use impairments and to identify the most likely causes of these problems. Depending on the magnitude of the effort expended and the clarity of the problems in the local area, it may also be possible to quantify the magnitude of stream use improvements with different levels of reduction of the causative agent. Once the causes and sources of the problems are identified, choices pertaining to improvement, or prevention measures in other areas, can be examined.

The following sections outline the concept of “weight-of-evidence” as a tool to assemble a large amount of data to help in obtaining needed information pertaining to environmental health. An example risk assessment is also provided to show how risks associated with exposures to humans can be examined.

DATA INTERPRETATION 611

EVALUATING BIOLOGICAL STREAM IMPAIRMENTS USING THE WEIGHT-OF-EVIDENCE APPROACH

The Process

The term “weight-of-evidence” (WOE) has been used frequently during the past several years in the environmental assessment arena. However, there is no clear definition or approach accepted, and approaches have varied from those that are crude and qualitative to very complex and quantitative. As discussed in Chapter 4, no one assessment approach is adequate for drawing conclusions on the quality of a waterway because of the associated uncertainties and weaknesses of each approach. Therefore, there is now widespread acceptance that multiple approaches (lines of evidence) are essential in order to reach reliable conclusions of whether a problem exists. Using the WOE approach, however, does not ensure that accurate conclusions will be obtained. It is critical that a well-designed assessment design be used (see Chapter 4) and that the key ecosystem components (biological, chemical, and physical) be characterized correctly, noting their associated uncertainties. The following discussion presents useful approaches for WOE evaluations.

One of the first WOE approaches to gain widespread attention was the “sediment quality triad” (Chapman et al. 1987). In this approach, sediment toxicity, indigenous biota, and sediment chemistry were characterized at each test site and normalized as a percentage of the reference (background) site condition. Results were presented graphically in an X-Y-Z axis type format. Comparing test site conditions to reference sites has long been used, but the primary contribution of the triad was to promote the notion that components must be assessed together. In stormwater assessments, the triad approach should be expanded to include the physical conditions (i.e., habitat), water and sediment conditions, and the associated temporal dynamics of each assessment component (Table 8.1).

Unfortunately, it is difficult to make quantitative evaluations of significant differences from this original “triad” approach. The comparisons between sites are particularly difficult at intermediate levels of contamination or if significant variability exists in the monitoring data. This “weight-ofevidence” approach can be evaluated using both parametric and nonparametric procedures to address the following study objectives: which stations are significantly different (impacted) relative to other stations?; how do the stations relate to each other?; and which parameters (monitoring components)

Table 8.1 0Summary of Key Weight-of-Evidence Components for Assessing Stormwater Effects on Receiving Waters

Component Media Priority Flow Level Difficultya

Benthic community Sediment High Low Low Fish community Water Medium Low Medium Toxicity

Lab-based Sediment Medium Low Low–Med. Water High Low and High Low–Med.

In situ-based Sediment High Low Low–Med. Water High Low and High Low–Med.

Bioaccumulation Benthic species Organism Medium Low Med.–High Fish species Organism Medium Low Med.–High In situ passive Water Low Low and High Med.–High

Chemistry (metals, Sediment High Low Med.–High organics, conventional Water High Low and High Med.–High physicochemistry)

Physical Flow Water High Low and High Low Habitat Whole stream High Low Low

a Difficulty rating considers both level of effort and cost to measure by typical approaches described in Appendices.


are significantly different (impacted) relative to other stations? Initial exploratory data analyses should be used to identify relationships between variables, identify and rank important variables, and identify weighting factors or redundant variables (i.e., responses mimic each other). These analyses may include correlation analyses, scatterplots, and other ordination tests. Results can also be ranked, whereby endpoint measures are averaged at each station and stations are then ranked by performance. Sample average ranks can be compared to a critical value to determine if significant differences exist between stations. Ordination procedures can be used to determine distances among stations and endpoints (e.g., multidimensional scaling). Scatterplots will show similarity of ranked groups and the magnitude of relationships among measured endpoints.

The Massachusetts Department of Environmental Protection formalized the WOE approach (Menzie et al. 1996) for relating measurement endpoints to assessment endpoints in ecological risk assessments. They identified three major components:

1. Weight assigned to each measurement endpoint: measurement endpoints (e.g., mortality, growth) may vary in the degree they relate to the assessment endpoints, or their quality, and may therefore be assigned differing levels of weight (i.e., importance).

2. Magnitude of response in the measurement endpoint: a greater weight is assigned to strong responses. 3. Concurrence among measurement endpoints: there tends to be greater confidence in findings that agree

with other lines of evidence. However, disagreement between components does not negate their validity or importance. For example, aquatic species have varying levels of sensitivity to different chemicals, or sampling may induce artifacts. Concordance of findings is more likely when very high levels of contamination are present, causing acute toxicity, as opposed to lower chronic toxicity exposures.

Numerical weighting values (e.g., 1 to 3 or 1 to 5) are assigned to elements of the process via professional judgment. This weighting of relative importance of the various tools has been done using Delphi techniques where a group of environmental professionals is surveyed. For example, each measurement endpoint (such as species population number) could be rated as high, medium, or low for three attributes (strength of relationship to an assessment endpoint, such as fish catch, data quality, and study design). These three attribute ratings may then be summed to get an overall measurement weight (of 1 to 3). The reliability of this best professional judgment approach is obviously related to the quality and comprehensive expertise of the survey group. After weighting values are assigned, measured responses are multiplied by their respective weights and summarized. The evidence showing the relationship between exposure to a stressor and a biological response (e.g., an assessment endpoint) is then assessed for risk. This leads the assessor to the most critical point of the assessment where the question is asked: what is the relationship between exposure to the stressor of concern (e.g., suspended solids, zinc, pesticides, stormwater) and adverse biological effects? The WOE process will help answer this question. While the WOE is the preferred approach, it is not without its shortcomings. Aside from not being a simple standardized protocol, the WOE is also not strictly quantitative, requiring best professional judgment. Statistically significant differences and relationships cannot readily be determined for the overall, integrated process. Certainty and accuracy are ensured via greater weight that is obtained through sound, comprehensive, integrated assessments. More importantly, as the WOE process is used in an area, it becomes “calibrated” through experience and observation and can become fine-tuned to better represent actual changes that may be occurring.

Benchmarks

In the process of interpreting exposure and effect relationships, there are a number of tools that can be used, ranging from “benchmarks,” or deterministic approaches, to probabilistic methods. These are discussed in the following sections. Benchmarks refer to concentrations or levels of physical and chemical parameters above which adverse biological effects may occur. These are often derived from large scientific databases linking biological responses with exposures to


Table 8.2 Categories of Biological Impairment Benchmarks

Regional or National Water/Sediment Quality Criteria or Standards State, Provincial, or Regional Water Quality Standards Biological Criteria Threshold (Toxicity) Effect Levels for Water, Sediment, or Tissues Hazard Quotients (Threshold Level or Site Concentration vs. No Effect Level, Reference or Background Site) Percentile Distributions Statistical Significance of Test vs. Control or Reference

compounds. Examples of commonly used benchmarks are listed in Table 8.2. Specific benchmarks/criteria for water and sediment criteria and biota are also discussed in Appendices B, C, and G. For each of these benchmark categories, there exists chemical specific benchmarks calculated by a variety of methods. These methods vary in the amount of biological effect (toxicity) information they include, ranging from only acute toxicity information on one species, to acute and chronic toxicity on many species. In addition, the toxicity information generated in these benchmarks ranges from a site-specific nature to being applicable to large geographical areas (such as north America). As with any assessment tool, each has associated uncertainties that should be recognized and considered by the assessor. The optimal approach is to use multiple benchmarks to better ascertain whether impairment exists.

The most important issue to remember when using benchmarks is that they are simply “benchmarks” to use in the chemical-physical data interpretation process. They do not unequivocally determine whether adverse effects are occurring. Often, these benchmarks do not include sitespecific biological effects data. In addition, the biological effect benchmarks may not be applicable to the conditions at your study site. For example, a suite of stressors may exist at your sites that interact to produce antagonistic or synergistic effects or conditions may alter the bioavailability of the chemical of concern. However, the use of multiple benchmarks that have been derived from large, scientifically valid, databases will assist in the weighting and data interpretation process.

The optimal method of establishing a relationship between biological effects and a site-specific parameter(s) is to thoroughly characterize exposure and effects. Benchmarks, unfortunately, only suggest that effects may be occurring if they are exceeded. If exceeded, they should at least be treated as “red flags,” emphasizing areas where additional investigation is warranted. They do not address spatial and temporal variability or site-specific interactions. This requires carefully designed biological and toxicity studies during low and high flow conditions (Chapters 4 and 6). In the absence of site-specific effects data, use of probability modeling is preferred if adequate site data exist for determining spatial and temporal exposure-effects interactions (see Ecological Risk Assessment section below).

Perhaps the best recognized and accepted benchmarks are the U.S. EPA’s National Ambient Water Quality Criteria (see Appendix G), which many states have adopted as their ambient water quality standards. The results of stormwater quality analyses have commonly been compared to water quality criteria in order to identify potentially toxic waters, and likely problematic pollutants. This has led to numerous problems with the interpretation of the data, especially concerning the “availability” of the toxicants to receiving water organisms and the exposure durations in receiving waters. The quality of stormwater, or of ambient waters immediately following high flow events, has been shown to be degraded in many studies with chemical concentrations that may exceed toxicity thresholds (e.g., Horner et al. 1994; Makepeace et al. 1995; Morrison et al. 1993; Waller et al. 1995). Stormwater toxicants are primarily associated with particulate fractions and are typically assumed to be “unavailable.” Typically short and intermittent runoff events can also not be easily compared to the “long” duration criteria or standards. Chemical analyses, without biological analyses, would have underestimated the severity of the problems because the water column quality varied rapidly, while the major problems were associated with sediment quality and effects on macroinvertebrates (Lenat and Eagleson 1981; Lenat et al. 1981).


The contradictions noted between in-stream biological effects and water quality criteria should not be surprising, given the assumptions used by the EPA:

1. Single acute and chronic average exposure period that does not account for pulse or repeated exposures for short time periods

2. Single bioavailability normalization factors (such as hardness) 3. Laboratory-derived toxicity values for surrogate species are protective of indigenous species 4. Effects derived from single chemical exposures in clean solutions where the toxicant is in the

dissolved form 5. Chemical exposures in the field based on limited grab sample analyses

To address magnitude and duration issues, the EPA developed the “Criterion Maximum Concentration” concept, with an exposure period assumption of 1 hour, and the “Criterion Continuous Concentration,” with an average period assumption of 4 days. Yet, these assumptions do not accurately describe most wet weather runoff exposures. Tests with pentachloroethane (Erickson et al. 1989, 1991) showed that with intermittent exposures, higher pulse concentrations were needed to affect growth, and when averaged over the entire test, effects were elicited at concentrations lower than when under constant exposure. The simplest toxicity model (with first-order, singlecompartment toxicokinetics and a fixed lethal threshold) could not completely describe the data. Erickson et al. (1989) concluded that kinetic models which predict mortality were reasonable; however, chronic toxicity effects were much more complicated, and no adequate models existed. Hickie et al. (1995) describe a one-compartment, first-order kinetics, pulse exposure model for residue-based toxicity of pentachlorophenol to P. promelas. Pulse exposures were of 2 min to 24 hours with durations of 2 to 24 hours, repeated 2 to 15 times. A comparison of three models (Cxt, Mancini, Breck 3 dimensional range repair) showed reasonable prediction of fish toxicity following 1 to 4 monochloramine pulses (2-h pulse, 22-h recovery). However, predictive capability decreased with greater than 4 pulses (Meyer et al. 1995). Beck et al. (1991) examined the transient nature of receiving water effects associated with stormwater, stressing the weaknesses associated with more typical steady-state approaches. They felt that there were still major misconceptions associated with modeling these effects.

Despite these limitations, water quality criteria and standards have been used effectively to identify potential stormwater problems and direct further assessment studies (see Chapter 6). Nonetheless, it is apparent that the use of water quality criteria to identify potential receiving water problems should be done with care. In many cases, the most direct comparison is made for concentrations of the soluble forms of the pollutants only and to use the short-term acute exposure criteria. This seems to be the most conservative approach, and if any measured pollutant exceeds this critical value, a problem pollutant is easily flagged. However, this approach is fraught with false negatives, as many chronic problems may still exist that are not recognized. As an example, numerous in-stream receiving water investigations (described in Chapter 3) have identified severe problems (indicated by lack of sensitive species) where the measured water quality met the criteria. Because the toxicants are strongly associated with particulates, secondary sediment contamination occurs that may be more important than water column conditions for aquatic life effects. In addition, habitat degradation caused by urbanization and agricultural activity (including highly fluctuating flows) are also likely responsible for many of the recognized receiving water problems. Finally, the irregular, but frequent, exposures of pollutant concentrations lower than the criteria may cause a greater problem than relatively constant, but higher, concentrations (see also Chapters 4 and 6). Therefore, direct comparisons of water quality criteria with monitored in-stream concentrations should be carefully conducted and used as adjuncts to direct in-stream biological use observations, plus evaluations of habitat and sediment quality. Human health criteria (such as pathogens for water-contact recreation and toxicants for drinking water supplies of fish/shellfish consumption) are more applicable to wet weather conditions and can be more directly used to flag potential problem pollutants.


Table 8.3 NURP Reported Median and 90th Percentile Event Mean Concentrations (EMC) (mg/L, unless otherwise noted) for Urban Runoff

Event to Event Median Urban Variability in 90th Percentile

Constituent Site EMC EMC (COV) Urban Site EMC

Suspended solids 100 1–2 300 BOD5 9 0.5–1.0 15 COD 65 0.5–1.0 140 Total P 0.33 0.5–1.0 0.70 Soluble P 0.12 0.5–1.0 0.21 TKN 1.5 0.5–1.0 3.3 NO2 + NO3 (as N) 0.68 0.5–1.0 1.8 Total copper (µg/L) 34 0.5–1.0 93 Total lead (µg/L) 144 0.5–1.0 350 Total zinc (µg/L) 160 0.5–1.0 500

From EPA (U.S. Environmental Protection Agency). Final Report for the Nationwide Urban Runoff Program. Water Planning Division, Washington, D.C. December 1983.

Appendix G presents a summary of the human health and aquatic health criteria for pollutants that commonly occur in urban runoff and receiving waters. Most of the criteria are expressed with a recommended exceedance frequency of 3 years. This is the EPA’s best scientific judgment of the average amount of time it will take an unstressed system to recover from a detrimental event in which exposure to the pollutant exceeds the criterion. A stressed system, for example, one in which several outfalls occur in a limited area, would be expected to require more time for recovery. Obviously, if criteria are exceeded for most rain events (such as can occur for bacteria and total recoverable heavy metals), then 3 years are not available for recovery before the next runoff event.

The discussions on the effects of the pollutants on aquatic life and human health presented in Appendix G are summarized from the U.S. EPA’s Quality Criteria for Water, 1986 (EPA 1986). The criteria were also reviewed using the EPA’s web page (http:/www.epa.gov) on the Internet for more recent changes. Some minor changes have been made since 1986 (chloride standards, for example, in 1988). Numeric criteria for heavy metals have been proposed as part of the states’ Compliance for Toxic Pollutants (for the nine states subject to EPA’s 1992 National Toxics Rule) as an interim rule. In most cases, only the short-term criteria are applicable for wet-weather receiving water conditions. Most runoff events last only a few hours; very few last for several days. However, degraded in-stream conditions can occur for several times the duration of the rain event itself. In addition, frequent exposures to concentrations less than the critical short-term criteria may result in significant problems that would not be predicted based on these criteria alone.

In some instances, acceptable stormwater concentration guidelines may be based on typical data as obtained during the Nationwide Urban Runoff Program (NURP) (EPA 1983). These data were almost solely represented by medium-density residential area runoff, with some data from other areas (such as shopping centers and light industrial areas). Useful benchmarks include the event mean concentrations, or EMC, (average of all observed concentrations) and the 90th percentile values of common parameters as measured during NURP (Table 8.3). The 90th percentile values are sometimes used as an upper limit for acceptable concentrations.

Ranking and Confirmatory Studies

If an adequate stormwater runoff study design is implemented (Chapter 4) and the weight-ofevidence process followed with the ensuing monitoring data, then sound decisions can be made. In reality, few comprehensive stormwater assessments look at all possible stressors and species of concern while characterizing the spatial and temporal dynamics of the system. Most environmental assessments are resource limited, requiring a tiered approach, where potential problem sites are identified, ranked, and then decisions made as to what further assessments are necessary. This is


Tier 1: Stress Demonstration

Tier 2: Stressor Class Identification

Tier 3: Stressor & Source Confirmation

Site Reconnaissance

Sample Design Issues

EffectsExposure reference sites vs. stressor gradient

.Water column

.Bioaccumulation -tissue design .PAHs -phototox testing.GW/SW interactions-piezometer design

.Surficial sediment

.Interface (sed/water)

.Pore water

Physicochemical Profiles

.Physical stressors(flow, temperature, suspended solids).Chemical stressors(PAHs, nonpolars, metals, ammonia ) classes .In Situ testing -Stressor Identification Evaluations (SIE).Laboratory testing -Toxicity Identification Evaluation (TIE); Phase 1

Compartment .Low flow .High flow .Seasonal .Diel

Event .1-30 days .H. azteca .D. magna.C. dubia .P. promelas .C. tentans .L. variegatus.etc.

Period Species

.Survival

.Labtox testing.Chemistry

.Habitat (QHEI).Retrospective studies

.Indigenous biota (structure/function indices, genetic profiling, fish DELTs, hyporheous)

.Growth .Reproduction .Tissue

Measurement Endpoints

Weight of Evidence

Figure 8.1 Example of a tiered weight-of-evidence approach used by Wright State University to evaluate stormwater runoff and aquatic ecosystem contamination.

particularly true for stormwater studies, where historical funding mechanisms do not exist and where the watershed-receiving water relationships are complex. It is important to be realistic in the expectations of initial screening studies. The goals should be to simply rank problem sites through the WOE approach (see above WOE discussion). Then, follow-up confirmatory (Tier 2) studies can focus on fewer sites, allowing for more quantitative characterization of the temporal dynamics and resulting effects of runoff events (Figure 8.1) (see also Chapter 4 example outline of a comprehensive runoff effect study).

For example, an approach used to identify stressors in aquatic ecosystems used by Wright State University is shown in Figure 8.1. During initial site reconnaissance, a determination is made as to whether three common sample design issues need to be incorporated: (1) Do pollutants (such as PCBs) that readily bioaccumulate likely occur? (2) Do PAHs likely occur? and/or, (3) Are there likely groundwater–surface water transition zones occurring in the area of contamination? If any of these three issues are present, then the typical Tier 1 sampling design may be modified to include: tissue residue or bioaccumulation testing, phototoxicity evaluations, and piezometer measures of groundwater movement (with concurrent chemical and toxicity testing of those compartments). The typical Tier 1 design will involve toxicity testing of two to four species which are exposed to three to four compartments (overlying water, sediment-water interface, surficial sediment, pore water) during low flow. At high flows, these same species are exposed to overlying waters. During their exposures, basic water quality measures are made, such as DO, conductivity, alkalinity, hardness, pH, temperature, turbidity (or TSS), and ammonia. If toxic effects are noted following these exposures, then Tier 2 testing may commence to better identify the type of stressor. Tier 2 testing may then require more in-depth chemical analyses and try to separate out stressors such as ammonia, metals, and nonpolar organics. Finally, in Tier 3, the focus can be to determine the significance of the dominant stressor(s) via a WOE approach.

The WOE process lends itself easily to ranking sites — particularly using broad categories such as high, medium, and low priority. For instance, this may separate sites that have acute toxicity and few pollution-sensitive benthic organisms from those with possible chronic toxicity and marginal benthic communities. The decision maker may then choose to immediately pursue installation of stormwater controls at the worst site, while conducting confirmatory studies at the marginal site to establish the extent and/or cause of the problem. Confirmatory studies are frequently necessary


to establish the: (1) dominant stressor(s); (2) exposure pathways/dynamics; (3) receptor organisms; (4) food web interactions; (5) environmental risk (human and ecological significance of effects); and (6) stressor sources. Confirmatory or Tier 2 studies are designed to answer very focused questions and use many of the same tools described for the more routine stormwater assessments. However, as the questions may be more focused, more specific and novel assessment techniques may be employed, such as DNA fingerprinting (RAPD PCR), toxicity identification evaluations (TIEs), or SPMDs (all described in Chapter 6). The environmental quality of many of our agricultural and urban waterways will also be less than pristine where anthropogenic influences are minimal. Therefore, the issue of whether significant ecological impairment exists will be more of a challenge in these human-dominated watersheds. The point of comparison for determinations of impairment should be an appropriate ecoregion reference or criteria, where manageable stressors have been removed (such as high temperature, erosion, pesticides, lack of riparian zone).

Comments Pertaining to Habitat Problems and Increases in Stream Flow

Habitat changes due to urbanization and agricultural activities are likely the cause for much of the degradation noted in biological conditions in streams. Appendix A outlines habitat evaluation schemes, while Chapter 6 also included descriptions on characterizing habitat. Understanding the effect that habitat has on stream biological uses is very important if these changes are to be minimized. This understanding needs to come from detailed local investigations, as our ability to predict habitat changes associated with stormwater discharges is rather poor. With site studies, some researchers have been able to recommend local guidelines to minimize habitat degradation.

MacRae (1997) found that stream bed and bank erosion is controlled by the frequency and duration of the mid-depth flows (generally occurring more often than once a year), not the bank-full condition (approximated by the 2-year event). During monitoring near Toronto, he found that the duration of the geomorphically significant predevelopment mid-bankfull flows increased by a factor of 4.2, after 34% of the basin had been urbanized, compared to before-development flow conditions. The channel had responded by increasing in cross-sectional area by as much as three times in some areas, and was still expanding. He also reported other studies that found channel cross-sectional areas began to enlarge after about 20 to 25% of the watershed was developed, corresponding to about a 5% impervious cover in the watershed. When the watersheds are completely developed, the channel enlargements were about five to seven times the original cross-sectional areas. Changes from stable stream bed conditions to unstable conditions appear to occur with basin imperviousness of about 10%, similar to the value reported previously for serious biological degradation. MacRae concluded that an effective criterion to protect stream stability (a major component of habitat protection) must address midbankfull events, especially by requiring similar durations and frequencies of stream power at these depths, compared to satisfactory reference conditions.

Much research on habitat changes and rehabilitation attempts in urban streams has occurred in the Seattle area of western Washington over the past 20 years. Sovern and Washington (1997) described the in-stream processes associated with urbanization in this area. The important factors that affect the direction and magnitude of changes in a steam’s physical characteristics due to urbanization include:

• The depths and widths of the dominant discharge channel will increase directly proportional to the water discharge. The width is also directly proportional to the sediment discharge. The channel width divided by the depth (the channel shape) is also directly related to sediment discharge.

• The channel gradient is inversely proportional to the water discharge rate and is directly proportional to the sediment discharge rate and the sediment grain size.

• The sinuosity of the stream is directly proportional to the stream’s valley gradient and is inversely proportional to the sediment discharge.

• Bedload transport is directly related to the stream power and the concentration of fine material and inversely proportional to the fall diameter of the bed material.


In their natural state, small streams in forested watersheds in western Washington have small low-flow channels (the aquatic habitat channel) with little meandering (Sovern and Washington 1997). The stream banks are nearly vertical because of clayey bank soils and heavy root structures, and the streams have numerous debris jams from fallen timber. The widths are also narrow, generally from 3 to 6 feet wide. Stable forested watersheds also support about 250 aquatic plant and animal species along the stream corridor. Pool/riffle habitat is dominant along streams having gradients less than about 2% slope, while pool/drop habitat is dominant along streams having gradients from 4 to 10%. The pools form behind large organic debris (LOD) or rocks. The salmon and trout in western Washington have evolved to take advantage of these stream characteristics. Sovern and Washington (1997) point out that less athletic fish species (such as chum and pink salmon) cannot utilize the steeper gradient, upper reaches of the streams. However, coho, steelhead, and cutthroat can use these upper areas.

Urbanization radically affects many of these natural stream characteristics. Pitt and Bissonnette (1984) reported that coho and cutthroat were affected by the increased nutrients and elevated temperatures of the urbanized streams in Bellevue, as studied by the University of Washington as part of the EPA’s NURP project (EPA 1983). These conditions were probably responsible for accelerated growth of the fry, which were observed to migrate to Puget Sound and the Pacific Ocean sooner than their counterparts in the control forested watershed, which was also studied. However, the degradation of sediments, from decreased particle sizes, adversely affected their spawning areas in streams that had become urbanized.

Sovern and Washington (1997) reported that in western Washington, frequent high flow rates can be 10 to 100 times the predevelopment flows in urbanized areas, but that the low flows in the urban streams are commonly lower than the predevelopment low flows. They have concluded that the effects of urbanization on western Washington streams are dramatic, in most cases permanently changing the stream hydrologic balance by: increasing the annual water volume in the stream, increasing the volume and rate of storm flows, decreasing the low flows during dry periods, and increasing the sediment and pollutant discharges from the watershed. With urbanization, the streams increase in cross-sectional area to accommodate these increased flows, and headwater downcutting occurs to decrease the channel gradient. The gradients of stable urban streams are often only about 1 to 2%, compared to 2 to 10% gradients in natural areas. These changes in width and the downcutting result in very different and changing stream conditions. The common pool/drop habitats are generally replaced by pool/riffle habitats, and the stream bed material is comprised of much finer material, for example. Along urban streams, fewer than 50 aquatic plant and animal species are usually found. Researchers have concluded that once urbanization begins, the effects on stream shape are not completely reversible. Developing and maintaining quality aquatic life habitat is possible under urban conditions, but it requires human intervention and it will not be the same as for forested watersheds.

Other Seattle area researchers have specifically examined the role that large woody debris (LWD) has in stabilizing the habitat in urban streams. Booth and Jackson (1997) found that LWD performs key functions in undisturbed streams that drain lowland forested watersheds in western Washington. These important functions include dissipation of the flow energy, channel bank and bed stabilization, sediment trapping, and pool formation. Urbanization typically results in the almost complete removal of this material. They point out that logs and other debris have long been removed from channels in urban areas for many reasons, especially because of their potential for blocking culverts or forming jams at bridges. Also, they may increase bank scour, and many residents favor “neat” stream bank areas (a lack of woody debris in and near the water and even with mowed grass to the water’s edge).

It is clear that stream hydraulics, sediment transport, and riparian vegetation dramatically affect habitat in streams. Water quality evaluations, by themselves, obviously do not consider these important factors. Evaluations of habitat conditions and effects of changing habitat on biological uses obviously require combinations of stream studies, modeling, and comparison with local


reference streams. The ability to predict habitat changes associated with urbanization, and the general success of habitat restoration efforts, is currently very poor. However, it is clear that detailed local investigations are critical and that habitat changes are likely one of the most important detrimental effects associated with urbanization.

EVALUATING HUMAN HEALTH IMPAIRMENTS USING A RISK ASSESSMENT APPROACH

The risk assessment paradigm is now well established in North America. The approach is basically the same for human health and ecological risk assessments (EPA 1989, 1998). The risk assessment paradigm is comprised of the following components: problem formulation, exposure and effects characterization followed by risk characterization, then the final risk management decisions. Risk assessment is a broad term which encompasses both risk characterization and risk management. The distinction between these two terms is an important one. The National Research Council’s 1983 report on risk assessment in the federal government distinguished between risk assessment and risk management.

Broader uses of the term [risk assessment] than ours also embrace analysis of perceived risks, comparisons of risks associated with different regulatory strategies, and occasionally analysis of the economic and social implications of regulatory decision functions that we assign to risk management. (EPA 1995)

The U.S. EPA has made the additional distinction of separating risk assessment from risk characterization. Risk characterization is the last step in risk assessment, is the starting point for risk managers, and is the foundation for regulatory decision making. The risk characterization identifies and highlights the noteworthy risk conclusions and related uncertainties (EPA 1995). The process described above is similar, but we have used different terminology. If the stormwater assessor is more comfortable using the EPA risk assessment approach, it can incorporate the guidelines of this handbook. The EPA guidance for conducting ecological risk assessments (ERAs) is quite general and does not provide specific methodologies and processes (EPA 1998). A number of good references (e.g., Suter 1993) exist that describe risk assessment approaches and considerations which are beyond the scope of this handbook. The two principal approaches for assessing adverse effects (hazard) in risk assessments are briefly described below.

Deterministic Approach

The simplest approach is the benchmark approach. This method (described above) basically ignores temporal exposure issues and focuses on point-in-time evaluations where threshold effect levels are compared to site contamination levels to ascertain risk. Many of the commonly used benchmarks (Appendix G) can be found in databases such the EPA’s Ambient Water Quality Database, state water quality standards, ECOTOX, and the U.S. Department of Energy’s Oak Ridge National Laboratory web site (http://www.hsrd.ornl.gov/ecorisk/ecorisk.html). This approach uses the quotient method for screening-level risk assessments. For compounds that bioaccumulate, it is easy to rearrange exposure equations involving uptake to back calculate ecotoxicity criteria for sediment, surface water, or soil (e.g., Pastorok et al. 1996).

Probabilistic Approach

A potentially more accurate and powerful assessment approach uses probability estimates to link likelihood of exposure with effects. This approach has been used frequently at Superfund sites,


looking at exposure pathway analysis and risk modeling to assess chemical risks to humans, and aquatic and terrestrial wildlife. Since food is a primary source of toxicants, food web models are important tools to describe potential ecosystem effects (Pastorok et al. 1996). The more advanced wildlife exposure models now contain three attributes: habitat spatial structure, food web complexity, and receptor behavior and physiology ranging from Tier 1 (steady-state, worst-case conservative) to Tier III (dynamic, stochastic). For assessments of aquatic stressor impacts, probabilistic assessments of pesticide effects have been conducted using the following steps (Solomon et al. 1996; World Wildlife Fund 1992):

1. Characterize sensitivity effects (select appropriate measurement endpoints and rank effect, e.g., EC50 or no observed effect levels, vs. concentration)

2. Characterize exposure (plot distribution of chemical concentrations vs. site vs. frequency of occurrence)

3. Risk characterization: compare exposure distribution with overlap of the sensitivity distribution, while considering uncertainty, confounding stressors, variables, and ecological relevance of the assessment

Example Risk Assessment for Human Exposure to Stormwater Pathogens

The following discussion, summarized from Meyland et al. (1998), describes waterborne pathogens in separate sewer overflow (SSO) discharges as an example of the risk assessment process applied to a wet-weather problem. SSOs are generally sanitary sewage discharges that occur at “relief” locations, resulting in untreated wastewater being discharged directly into receiving waters.

Hazard Identification

The first step in a risk assessment, hazard identification, can be examined by gathering information regarding waterborne disease outbreaks. The agent that causes disease could be chemical, physical, or biological. However, in this case we will focus on biological causes, or infectious agents, i.e., pathogenic microorganisms. Table 8.4 shows the agents that have caused waterborne disease outbreaks in the United States, from 1971 to 1990. Notice that the vast majority of known agents are microorganisms. Table 8.5 shows additional data compiled from waterborne disease outbreaks. This table shows the agent associated with the disease.

The Centers for Disease Control (CDC) keep detailed records regarding notifiable, or reportable, diseases. There are legal requirements for reporting cases of these diseases. This list of notifiable diseases includes cryptosporidiosis. The fact that this disease is notifiable means that it is recognized as being extremely hazardous. As of mid-April 1998, there were 520 cases of cryptosporidiosis (not notifiable in all 50 states) (CDC 1998).

Dose–Response

The concept of dose–response, the second step in a risk assessment, is critical. Briefly, dose–response describes a relationship between a given level of contaminant and the biological response induced. This relationship is usually incremental; i.e., increase in the dose causes an increase in the response. In this particular case, the dose is the number of pathogenic microorganisms that the human subject is exposed to (through ingestion, swimming, wading, etc.), and the response is the level of infection. Generally, there is a minimum infective dose threshold that must be reached in order to infect a given individual. Once an individual has been infected, there are increasing degrees of infection severity. A subclinical infection describes the case where the pathogen produces a detectable immune response or organisms may be found growing in the human host, but the subject exhibits no clinical signs or symptoms, e.g., diarrhea, vomiting, etc. A clinical infection


Table 8.4 Causative Agents of Waterborne Disease Outbreaks, 1971 to 1990

Outbreaks Illness Number Percentage Number Percentage of Cases of Total of Cases of Total

Gastroenteritis (unknown cause)

Giardiasis Chemical poisoning Shigellosis Viral gastroenteritis Hepatitis A Salmonellosis Camplylobacterosis Typhoid fever Yersiniosis Cryptosporidiosis Chronic gastroenteritis Toxigenic E. coli Cholera Dermatitis Amebiasis

293 49.66 67,367 47.60 110 18.64 26,531 18.75 55 9.32 3877 2.74 40 6.78 8806 6.22 27 4.58 12,699 8.97 25 4.24 762 0.54 12 2.03 1370 0.97 12 2.03 5233 3.70

5 0.85 282 0.20 2 0.34 103 0.07 2 0.34 13,117 9.27 1 0.17 72 0.05 2 0.34 1243 0.88 1 0.17 17 0.01 1 0.17 31 0.02 1 0.17 4 0.00

Cyanobacteria-like bodies 1 0.17 21 0.01 Total 590 100 141,535 100

Data from Committee on Groundwater Recharge, NRC (National Research Council), National Academy of Science. Ground Water Recharge Using Waters of Impaired Quality. National Academy Press, Washington, D.C. 284 pp. 1994.

Table 8.5 Waterborne Disease Outbreaks Due to Microorganismsa

Disease Agent Outbreaksb (%) Casesc (%)

Bacteria Typhoid fever Salmonella typhi Shigellosis Shigella spp. Salmonellosis Salmonella paratyphi and other Salmonella species Gastroenteritis Escherichia coli

Campylobacter spp. Viruses

Infectious hepatitis Hepatitis A virus Diarrhea Norwalk virus

Protozoa Giardiasis Giardia lamblia Cryptosporidiosisd Cryptosporidium parvum

Unknown etiology Gastroenteritis

10 0.1 9 2.6 3 3.5 0.3 0.7 0.3 0.7

11 0.5 1.5 0.6

7 3.8 0.2 71

57 16.7

a Compiled from data provided by the Centers for Disease Control, Atlanta, GA. b Of more than 650 outbreaks in recent decades. c Of 520,000 cases over the same period. d A single outbreak of cryptosporidiosis in 1993 caused illness in 370,000 individuals from Milwaukee, WI. This

is the largest single recorded outbreak of a waterborne disease in history.

Data from Madigan, M.T., J.M. Martinko, and J. Parker. Brock Biology of Microorganisms, 8th ed. Prentice-Hall, Upper Saddle River, NJ. 1997.

refers to the condition in which there are clinical signs and symptoms present. In layman’s terms, one would refer to a person with a clinical infection as being “ill.” The most severe response to infection would be death, i.e., a fatality. Therefore, one usually refers to the MID50, that is, the minimum infective dose that will cause subclinical infection in 50% of people exposed to that number of pathogens. The minimal infective dose (MID) varies widely with the type of pathogen, as shown in Table 8.6 (Bitton 1994). Of those infected, a percentage will show clinical signs; this


Table 8.6 0Minimal Infective Doses for Some Pathogens and Parasites

Organism Minimal infective Dose

Salmonella spp. 104 to 107

Shigella spp. 101 to 102

Escherichia coli 106 to 108

Vibrio cholerae 103

Giardia lamblia 101 to 102 cysts Cryptosporidium 101 cysts Entamoeba coli 101 cysts Ascaris 1–10 eggs Hepatitis A virus 1–10 PFU

Data from Bitton, G. Wastewater Microbiology. John Wiley & Sons, Inc. New York. 1994.

Table 8.7 0Values Used to Calculate Risks of Infection, Illness, and Mortality from Selected Enteric Microorganisms

Probability of Infection from Exposure to One Ratio of Clinical Illness Secondary Organism (per million) to Infection (%) Mortality Rate (%) Spread (%)

Campylobacter Salmonella typhi Shigella Vibrio cholerae Coxsackieviruses Echoviruses Hepatitis A virus Norwalk virus Poliovirus 1 Poliovirus 3 Rotavirus Giardia lamblia

7000 380

1000 7

5–96 0.12–0.94 76 17,000 50 0.27–0.29 40

75 0.6 78 0.0001 30

14,900 0.1–1 0.9 90 31,000

310,000 28–60 0.01–0.12 19,800

Data from Committee on Groundwater Recharge, NRC (National Research Council), National Academy of Science. Ground Water Recharge Using Waters of Impaired Quality. ISBN 0-309-05142-8. National Academy Press, Washington, D.C. 284 pp. 1994.

is referred to as the ratio of clinical illness to infection. In addition, a percentage of those infections will result in fatalities; this is referred to as the case fatality rate. Table 8.7 shows example values for these various levels of response to infection.

Notice that higher probabilities, rates, or percentages correspond to pathogens with higher virulence. For example, if 1 million people are exposed to one rotavirus each, then 310,000 may be infected. In contrast, if 1 million people are exposed to one Vibrio cholerae bacterium each, only seven may be infected. In general, viral pathogens are much more virulent than bacterial pathogens.

Table 8.8 shows another example of data that can be obtained from published studies. These data show, for instance, that once infected by Salmonella bacteria, approximately 41% will exhibit clinical infection. In addition, Cryptosporidium infection results in a 71% clinical infection frequency.

Another study (DuPont et al. 1995) published results pertaining to infection rates from the oral introduction of Cryptosporidium oocysts into healthy volunteers. Various doses of oocysts, from 30 to 1 million, were given to volunteers in gelatin capsules, and these subjects were followed up to record the incidence of infection. Table 8.9 gives these results. A linear regression analysis of the data yielded a correlation coefficient of 0.983 and an infectious dose of 50 of 132 oocysts. This is an excellent example of the dose–response relationship, as increasing doses of oocysts caused increasing rates of infection.


Table 8.8 0Ratio of Clinical to Subclinical Infections and Case Fatality Rates for Enteric Microorganisms

Microorganism Frequency of Clinical Illness (%) Case:Fatality Rate (%)

Viruses Hepatitis (adults) 75 0.6 Rotavirus 25–60 0.01 Astrovirus (adults) 12.5 0.12 Coxsackie A16 50 0.59–0.94 Coxsackie B 5–96

Bacteria Salmonella 41 0.1 Shigella 46 0.2

Protozoan parasites Giardia 50–67 Cryptosporidium 71

Data from Gerba, C.P., J.B. Rose, C.N. Hass, and K.D. Crabtree. Waterborne rotavirus: a risk assessment. Water Research, Vol. 30, No 12, pp. 2929–2940. Dec. 1996.

Table 8.9 0Rate of Infection, Enteric Symptoms, and Clinical Cryptosporidiosis, According to the Intended Dose of Oocysts

Intended Dose No. of Enteric of Oocysts Subjects Infection Symptoms Cryptosporidiosis

Number (percent) 30 5 1 (20) 0 0

100 8 3 (37.5) 3 (37.5) 3 (37.5) 300 3 2 (66.7) 0 0 500 6 5 (83.3) 3 (50) 2 (33.3)

>1000 7 7 (100) 5 (71.4) 2 (28.6) Total 29 18 11 7

Data from DuPont, H.L., C.L. Chappell, C.R. Sterling, P.C. Okhuysen, J.B. Rose, and W. Jakubowski. The infectivity of Cryptosporidium parvum in healthy volunteers. N. End. J. Med., Vol. 332, No. 13, pp. 855–859. 1995.

Exposure Assessment

The exposure assessment is the third step in a risk assessment. Several factors contribute to whether or not contact with a particular pathogen may cause disease. Among these factors are virulence, mode of transmission, portal of entry, and host susceptibility. Virulence is defined as a particular organism’s ability to cause disease in humans and is related to the dose of infectious agent necessary for host infection and causing disease (Bitton 1994). The mode of transmission is the particular method in which the organism is transported from the reservoir of pathogens (such as a contaminated outfall) to the host, i.e., person-to-person, waterborne, or foodborne.

The research conducted by Meyland et al. (1998) concentrated on the waterborne transmission route of SSOs to exposed individuals, but exposure assessment can also be evaluated based on portal of entry. The portal of entry is dictated by the mechanism of contact; examples or entry portals are access through the gastrointestinal tract, respiratory tract, or skin. Host susceptibility is dependent upon resistance to infectious agents, which consists of the roles of the immune system and nonspecific factors (Bitton 1994). Immunity can be both natural (genetic) and acquired from previous contact with the pathogen.

There are many documented examples of waterborne transmission of pathogenic microorganisms. Recently, in the United States, there has been widespread concern about Cryptosporidium contamination of water supplies. This is an example of waterborne transmission via a contaminated drinking water supply. Table 8.10 summarizes the available information regarding Cryptosporidium outbreaks in the United States. Outbreaks caused by this organism are a


Table 8.10 0Cryptosporidium Outbreaks: Affected Populations and Characteristics of the Raw Water Supply

Estimated No. of People Affected Raw Water Suspected Sources of

County, State (City) Date (Confirmed Cases)a Source Contamination

Bexar County, TX (Braun Station)

Bernalillo County, NM (Albuquerque)

Carroll County, GA (Carrollton)

Berks County, PA

Jackson County, OR (Talent and Medford)

Milwaukee County, WI (Milwaukee)

Yakima County, WA

Cook County, MN (Grand Marais)

Clark County, NV (Las Vegas)

Walla Walla County, WA Alachua County, FL

May–Jul 1984

Jul–Oct 1986

Jan–Feb 1987

Aug 1991

Jan–Jun 1992

Jan–Apr 1993

Apr 1993

Aug 1993

Jan–Apr 1994

Aug–Oct 1994 Jul 1995

2,000 (47) Well Raw sewageb

(78) Surface water Surface runoff from

13,000 River

551 Well

15,000 Spring/river

403,000 Lake

7 (3) Well

27 (5) Lake

(78)c Lake

86 (15) Well (72) N/A

livestock grazing areas Raw sewage and runoff from cattle grazing areas

Septic tank effluent, nearby creek

Surface water, treated wastewater,b or runoff from agricultural areas

Cattle wastes, slaughterhouse wastes, and sewage carried by tributary rivers

Infiltration of runoff from cattle, sheep, or elk grazing areas

Backflow of sewage or septic tank effluent into distribution, raw water inlet lines, or both

Treated wastewater, sewage from boats

Treated wastewaterb

Backflow of contaminated water

a Estimates are based on epidemiologic studies; confirmed cases correspond to patients whose stool samples tested positive for Cryptosporidium.

b Strong evidence to support effect of wastewater. c 103 laboratory-confirmed cases were associated with the outbreak; 78 of these were documented during the

epidemiologic study period.

Data from Solo-Gabriele, H. and S. Neumeister. US outbreaks of cryptosporidiosis. Journal American Water Works Association, Vol. 88, No. 9, pp. 76–86. Sept. 1996.

significant health threat (more than 400,000 people were infected during the 1993 Milwaukee outbreak). Moreover, notice that the suspected source of contamination is likely to be sewage. In fact, municipal wastewater was implicated as the source in roughly half of the outbreaks (Solo-Gabriele and Neumeister 1996). The remaining outbreaks were likely caused by contaminated agricultural runoff.

Another important mode of transmission is water-contact recreation. This type of transmission is usually associated with swimming beach exposures. Many studies have shown an association between illness and swimming near stormwater, SSO, or CSO (combined sewer overflow) outfalls. In general, most of these studies found an increased risk of illness resulting from swimming in waters that contained fecal contamination indicators or pathogenic microorganisms. The SMBRP (Santa Monica Bay Restoration Project) Study is unique in that it found a distance-dependent association between contamination sources and health effects. In this study, there was a higher rate of enteric illness in swimmers who swam within 400 ft of a stormwater outfall than in those who swam more than 400 ft away. In many urban receiving waters, children frequently play in and near potentially contaminated small streams and creeks, well away from designated swimming beaches. In most cases, this exposure route is not considered, because this is not a designated use of the water. The most important pathogens contained in stormwater are likely from sewage contamination.


Table 8.11 Sensitive Populations in the United States

Population Individuals Year

Pregnancies 5,657,900 1989 Neonates 4,002,000 1989 Elderly people (over 65) 29,400,000 1989 Residences in nursing homes or related care facilities 1,553,000 1986 Cancer patients (non-hospitalized) 2,411,000 1986 Transplant organ patients (1981–1989) 110,270 1981–1989 AIDS patients 142,000 1981–1990

Data from Gerba, C.P., J.B. Rose, C.N. Hass, and K.D. Crabtree. Waterborne rotavirus: a risk assessment. Water Research, Vol. 30, No 12, pp. 2929–2940. Dec. 1996.

Pitt et al. (1993) conducted a study of inappropriate pollutant entries into storm drainage systems in which many illegal sanitary sewer connections to storm drain systems were commonly found.

Another possible exposure route is through the consumption of contaminated fish or shellfish. One example of this type of outbreak occurred in Louisiana in 1993 (Kohn 1995). This outbreak was caused by contaminated oysters that were consumed raw. The agent implicated in this outbreak was Norwalk virus, which causes gastroenteritis. Seventy of the 84 people (83% infection rate) who ate the raw oysters became ill. The epidemiologic investigation found that this outbreak was probably caused by overboard sewage disposal by harvesters near the oyster bed.

An additional consideration that one must account for when assessing the adverse health effects of contact with pathogenic microorganisms is that certain individuals within the population are at higher risk for serious infections. Individuals who are at higher risk are the very young, elderly people, pregnant individuals, and immunocompromised (organ transplants, cancer, and AIDS) patients (Gerba et al. 1996). This collective group represents almost 20% of the current U.S. population (Table 8.11). In addition, elderly and immunocompromised people are an increasing segment of the population.

Calculation of Risk

The fourth step in a risk assessment is the calculation of risk for a specific situation. As indicated in the first three steps, it is possible to identify which components pose a risk for a specific activity, to estimate the doses necessary to cause different responses, and to conduct the exposure assessment. For an urban receiving water study, it is possible to consider an important, but commonly overlooked, scenario: exposure to pathogens by children who are wading in an urban stream. The list of potential pathogens may be lengthy, depending on the likelihood of sanitary sewage contamination. Even with separate stormwater, numerous pathogens are still commonly present (such as Pseudomonas aeruginosa and Shigella). Sanitary sewage contamination lengthens the list considerably, as shown on the above lists. The dose–response curves can be examined for each likely pathogen and route of exposure (such as water contact for P. aeruginosa, or ingestion for Giardia and Cryptosporidium). The difficulty is estimating the magnitude of discharge for each pathogen and its fate after the discharge to the likely point of contact. Exposure duration, or ingestion quantity, of the water also needs to be estimated, along with likely time between discharge and exposure. This is especially important for wading exposures, for example, because water-contact recreation is unlikely during a runoff event but can certainly occur soon after a rain has ended. However, exposure to pathogens when wading is likely to be made more severe through stirring up contaminated sediments. Bacteria have been shown to survive for long periods in stream sediments in areas of deposition (pools, backwaters, or behind small dams), many of which attract children because of the deeper water (Burton et al. 1987; Burton 1982). Therefore, it would be necessary to predict areas where the particulates, with which pathogens are associated, are likely to accumulate.


The local monitoring program must therefore consider characterization studies of the pathogens in the discharge, along with the hydraulic characteristics of the receiving water that would affect transport and fate of the pathogens. This information is used in a receiving water model to predict the conditions (concentrations in the water) at the time and location of exposure. Appendix H summarizes representative receiving water models that may be applicable for certain fate and route pathways. HSPF, one of the more complex urban stream models, contains many options, but requires a great deal of monitoring data for calibration and verification. Other, simpler models, such as WASP5, may be suitable for this task. The predicted conditions at the likely sites of exposure can be used, along with assumptions pertaining to exposure duration. This information is then compared to the dose–response information to estimate the likelihood of contacting disease.

IDENTIFYING AND PRIORITIZING CRITICAL STORMWATER SOURCES

An important goal of a receiving water study is to learn enough about local problems to be able to reduce them in the future, and to prevent new problems from occurring. The tools and techniques presented in this book enable local watershed managers to obtain a good understanding of their local problems and their likely causes.

After receiving water problems (beneficial use impairments) are described, the likely causes for these problems must be identified. Many of the problems are directly associated with measured parameters in excessive quantities and are therefore inherently obvious (such as high bacteria concentrations interfering with water-contact recreation and fish consumption; high flows and debris blockages affecting drainage; trash or odors affecting noncontact recreation; and high toxicant concentrations affecting water supplies). The most difficult task is identifying the possible causes for declining aquatic life conditions, which can be associated with numerous causes, including habitat destruction, high/low flows, inappropriate discharges, polluted sediment, and water quality. The weight-of-evidence approach, described above, has therefore become a useful framework to understand possible cause-and-effect relationships for biological resources.

Once the critical pollutants (or flow conditions) are identified, it is more straightforward to identify the likely sources in the watershed contributing these problem constituents. Classical approaches include watershed modeling to develop mass balances for targeted pollutants. The following discussion describes this approach, along with a case study describing how the Wisconsin Nonpoint Source Program has integrated field monitoring, data analysis, watershed modeling, and the development of stormwater controls to reduce future receiving water problems. It is also important to recognize additional potential sources of contaminants in a watershed that are not easily addressed by most watershed models. The most important include snowmelt and inappropriate sources. These two sources can be much greater contributors of some critical pollutants than warm-weather runoff alone. Field investigations to locate and quantify inappropriate sources (sanitary sewage and discharges from small industries and commercial establishments are the most important) were described in Chapter 6 and must be conducted along with conventional watershed modeling activities. Snowmelt contributions should also be quantified for comparison with the more traditional sources.

Sources of Urban Stormwater Contaminants

Urban runoff is comprised of many separate source area flow components that are combined within the drainage area and at the outfall before entering the receiving water. It may be adequate to consider the combined outfall conditions alone when evaluating the long-term, areawide effects of many separate outfall discharges to a receiving water. However, if better predictions of outfall characteristics (or the effects of source area controls) are needed, the separate source area components


must be characterized. The discharge at the outfall is made up of a mixture of contributions from different source areas. The “mix” depends on the characteristics of the drainage area and the specific rain event. The effectiveness of source area controls is therefore highly site and storm specific.

Various urban source areas contribute different quantities of runoff and pollutants, depending on their specific characteristics. Impervious source areas may contribute most of the runoff during small rain events. Examples of these source areas include paved parking lots, streets, driveways, roofs, and sidewalks. Pervious source areas become important contributors for larger rain events. These pervious source areas include gardens, lawns, bare ground, unpaved parking areas and driveways, and undeveloped areas. The relative importance of each individual source is a function of their areas, their pollutant washoff potentials, and the rain characteristics.

The washoff of debris and soil during a rain is dependent on the energy of the rain and the properties of the material. Pollutants are also removed from source areas by winds, litter pickup, or other cleanup activities. The runoff and pollutants from the source areas flow directly into the drainage system, onto impervious areas that are directly connected to the drainage system, or onto pervious areas that will attenuate some of the flows and pollutants before they discharge to the drainage system.

Sources of pollutants on paved areas include on-site particulate storage that cannot be removed by the usual processes e.g., rain, wind, street cleaning, etc. Atmospheric deposition, deposition from activities on these paved surfaces (auto traffic, material storage, etc.), and the erosion of material from upland areas that discharge flows directly onto these areas are the major sources of pollutants to the paved areas. Pervious areas contribute pollutants mainly through erosion processes where the rain energy dislodges soil from between plants. The runoff from these source areas enters the storm drainage system where sedimentation in catchbasins or in the sewerage may affect their ultimate discharge to the outfall. In-stream physical, biological, and chemical processes affect the pollutants after they are discharged to the ultimate receiving water.

It is important to know when the different source areas become “active” (when runoff initiates from the area, carrying pollutants to the drainage system). If pervious source areas are not contributing runoff or pollutants, the prediction of urban runoff quality is much simplified. The mechanisms of washoff, and delivery yields of runoff and pollutants from paved areas, are much better known than from pervious urban areas (Novotny and Chesters 1981). In many cases, pervious areas are not active runoff contributors except during rain events greater than at least 5 or 10 mm. For smaller rain depths, almost all of the runoff and pollutants originate from impervious surfaces (Pitt 1987). However, in many urban areas, pervious areas may contribute the majority of the runoff, and some pollutants, when rain depths are greater than about 20 mm. The actual importance of the different source areas is highly dependent on the specific land use and rainfall patterns. Obviously, in areas having relatively low-density development, especially where moderate- and large-sized rains occur frequently (such as in the Southeast), pervious areas typically dominate outfall discharges. In contrast, in areas having significant paved areas, especially where most rains are relatively small (such as in the arid West), the impervious areas dominate outfall discharges. The effectiveness of different source controls is therefore quite different for different land uses and climatic patterns.

If the number of events exceeding a water quality objective is important, then the small rain events are of most concern. Stormwater runoff typically exceeds some water quality standards for practically every rain event (especially for bacteria and some heavy metals). In the upper Midwest, the median rain depth is about 6 mm, while in the Southeast, the median rain depth is about twice this depth. In most urban areas, no runoff is observed until the rain depth exceeds 2 or 3 mm. For these small rain depths and for most urban land uses, directly connected paved areas usually contribute most of the runoff and pollutants. However, if annual mass discharges are more important, e.g., for long-term effects, then the moderate rains are more important. Rains from about 10 to 50 mm produce most of the annual runoff volume in many areas of the United States. Runoff from both impervious and pervious areas can be very important for these rains. The largest rains (greater than about 100 mm) are relatively rare and do not contribute significant amounts of runoff pollutants


during normal years, but are very important for drainage design. The specific source areas that are most important (and controllable) for these different conditions vary widely.

Case Study: Wisconsin Nonpoint Source Program in Urban Areas

The urban stormwater evaluation methodology used by the Nonpoint Source Program at the Wisconsin Department of Natural Resources was developed to supplement the extensive ongoing rural aspects of the statewide watershed planning program (Pitt 1986). This comprehensive urban methodology includes:

• Evaluation of receiving water goals, based on established water use objectives and through meetings of citizen groups and technical experts

• Identification of current problems and sources of problem pollutants and flows through monitoring and modeling

• Identification and evaluation of suitable source area and outfall treatment options, including the development of model ordinances for construction site erosion control and stormwater management, plus the development of design manuals for constructing controls

• Demonstration projects evaluating alternative controls • Receiving water evaluations to confirm or to modify recommendations

An important element of this methodology was to extensively modify SLAMM, the Source Loading and Management Model (Pitt and Voorhees 1995; www.winslamm.com). This model was developed to identify sources of pollutants in urban areas and to evaluate many alternative stormwater control programs. This methodology is generally used in the development of the watershed plans and to determine suitable cost-sharing aspects of the Wisconsin Nonpoint Source Program. The Milwaukee area was the first urban watershed planning effort to use this methodology. Milwaukee has a great deal of stormwater quality information that was used in this planning and implementation effort.

Stormwater quality management in the Milwaukee area was initiated as part of the Wisconsin Priority Watershed Program. This program was developed in 1978 to help combat both urban and rural nonpoint sources of pollution (EPA 1990a, 1991). This program is one of the oldest in the nation funding nonpoint pollution abatement. An important element of this program is retrofitting control practices in both rural and urban areas. The program was initially heavily involved in rural areas, with technical assistance from the NRCS. A unique aspect of the program is that it is implemented on a watershed, and not on a political jurisdiction basis. Of the state’s 330 watersheds, 130 (mostly located in the southern part of Wisconsin) will likely require comprehensive management activities to control nonpoint pollutants. A 25-year plan was developed in 1982 which would require the start-up of about eight or nine new watershed abatement efforts per year. The watershed plans are prepared by the state with cooperation and reviews by local government agencies. They contain detailed analyses of the water resources objectives (existing and desired beneficial uses, including the problems and threats to these uses), the critical sources of problem pollutants, and the control practices that can be applied within each watershed. The plans also include implementation schedules and budgets to meet the pollution reduction objectives.

Each plan requires 1 year to prepare, including the necessary fieldwork. Various field inventory activities are needed to prepare the plans, including aquatic biology and habitat surveys to identify existing and potential fishery uses, stream bank surveys to identify the nature and magnitude of stream bank erosion problems and to help design needed controls, field and barnyard surveys to supply information needed to estimate and rank their pollution potentials and to design farm control practices, and urban surveys needed to evaluate urban runoff pollution potential and its control.

Urban planning was initiated in 1983 in the Milwaukee and Madison areas, with other urban areas of the state following. The urban practices eligible for cost sharing identified in these plans have included stream bank protection, detention basins, and infiltration devices for existing urbanized areas. Construction site erosion controls are also usually required as a condition for a grant


agreement in an urban area, but they are not eligible for state cost sharing. About $3 to $5 million per year will be used by the nonpoint source program over a 20-year period in controlling urban runoff. An outcome of the Milwaukee River South Watershed plan included goals for reducing urban stormwater discharges (D’Antuono 1998). These goals were 50% reductions for suspended solids and heavy metals, and 50 to 70% reductions for phosphorus.

Detailed studies on toxicant sources, effects, and controls have also been conducted in Milwaukee, including a study conducted in the heavily urbanized Lincoln Creek (having a 19-mi2 watershed and being 9 mi long). A seven-tiered indicator program, incorporating many physical, chemical, and biological tests, was simultaneously conducted which identified long-term toxicity problems, likely associated with resuspended contaminated sediments having high levels of organic compounds (Claytor 1996). It was found that discharges of these fine sediments could be significantly reduced through the use of well-designed and maintained wet detention basins. The in-stream toxicity monitoring methods developed and used during the Lincoln Creek study can be used by other municipalities to answer the following basic questions:

• Are toxic conditions present? • What is causing the toxicity? • How much is too much urbanization? • Can stormwater controls reduce these problems?

The benefits of stormwater controls have also been evaluated in Milwaukee, especially grass swales, wet detention ponds, and underground devices for critical source areas. The Southeastern Wisconsin Regional Planning Commission also prepared a comprehensive report documenting costs associated with construction site erosion and stormwater control.

Use of SLAMM to Identify Pollutant Sources and to Evaluate Control Programs

A logical approach to stormwater management requires knowledge of the problems to be solved, the sources of the problem pollutants, and the effectiveness of stormwater management practices that can control the problem pollutants at their sources and at outfalls. The Source Loading and Management Model (SLAMM) is designed to provide information on these last two aspects of this approach. The first versions of SLAMM were developed by Pitt in the mid-1970s to help evaluate the results from early EPA stormwater projects (Pitt and Voorhees 1995). Further information on SLAMM, especially its integration with GIS systems, is included in Appendix H and at www.winslamm.com.

The development of SLAMM began in the mid-1970s, primarily as a data reduction tool for use in early street cleaning and pollutant source identification projects sponsored by the EPA’s Storm and Combined Sewer Pollution Control Program (Pitt 1979, 1984; Pitt and Bozeman 1982). Additional information contained in SLAMM was obtained during the EPA’s Nationwide Urban Runoff Program (NURP) (EPA 1983), especially the early Alameda County, CA (Pitt and Shawley 1982), and the Bellevue, WA (Pitt and Bissonnette 1984) projects. The completion of the model was made possible by the remainder of the NURP projects and additional field studies and programming support sponsored by the Ontario Ministry of the Environment (Pitt and McLean 1986), the Wisconsin Department of Natural Resources (Pitt 1986), and Region V of the U.S. Environmental Protection Agency. Early users of SLAMM included the Ontario Ministry of the Environment’s Toronto Area Watershed Management Strategy (TAWMS) study (Pitt and McLean 1986) and the Wisconsin Department of Natural Resources’ Priority Watershed Program (Pitt 1986). SLAMM can now be effectively used as a tool to enable watershed planners to obtain a better understanding of the effectiveness of different control practice programs.

Some of the major users of SLAMM have been associated with the Nonpoint Source Pollution Control Program of the Wisconsin Department of Natural Resources, where SLAMM has been used for a number of years to support their extensive urban stormwater planning and cost-sharing program (Thum et al. 1990; Kim et al. 1993a,b; Ventura and Kim 1993; Bachhuber 1996; Banner-


man et al. 1996; Haubner and Joeres 1996; Legg et al. 1996). Many of these applications have included the integrated use of SLAMM with GIS models, as illustrated in Appendix H.

SLAMM was developed primarily as a planning-level tool to generate information needed for planning-level decisions, while not generating or requiring superfluous information. Its primary capabilities include predicting flow and pollutant discharges that reflect a broad variety of development conditions and the use of many combinations of common urban runoff control practices. Control practices evaluated by SLAMM include detention ponds, infiltration devices, porous pavements, grass swales, catchbasin cleaning, and street cleaning. These controls can be evaluated in many combinations and at many source areas as well as the outfall location. SLAMM also predicts the relative contributions of different source areas (roofs, streets, parking areas, landscaped areas, undeveloped areas, etc.) for each land use investigated. As an aid in designing urban drainage systems, SLAMM also calculates correct NRCS curve numbers that reflect specific development and control characteristics. These curve numbers can then be used in conjunction with available urban drainage

GR

AS

S S

TR

IP

procedures to reflect the water quantity reduction benefits of stormwater quality controls. S

IDE

WA

LK

SLAMM is normally used to predict source area contributions and outfall discharges. However, it has been used in conjunction with a receiving water model (HSPF) to examine the ultimate

FR

ON

TYA

RD

receiving water effects of urban runoff (Ontario 1986), and has been recently been modified to be integrated with SWMM (Pitt et al. 1999) to more accurately consider the joint benefits of source

DR

IVE

WA

YS

area controls on drainage design. The following example illustrates how SLAMM is used to identify the most important source areas and how to select the most appropriate control programs.

WA

LK

WA

YS

The areas of the different surfaces in each land use are very important for SLAMM evaluations. B

AC

KYA

RD

Figure 8.2 is an example showing the areas of different surfaces for a medium-density residential area in Milwaukee. As shown in this example, streets make up between 10 and 20% of the total

PO

OL

S

area, while landscaped areas can make up about half of the drainage area. The variation of these different surfaces can be very large within a designated area. The analysis of many candidate areas

DE

CK

S/S

HE

DS

might therefore be necessary to understand how effective or how consistent the model results might be for a general land use classification.

One of the first problems in evaluating an urban area for stormwater controls is the need to understand where the pollutants of concern are originating under different rain conditions.

6000

5000

4000

3000

2000

1000

0

m2 /

ha

ST

RE

ET

RO

OF

TOP

S

Figure 8.2 Source areas — Milwaukee medium-density residential areas (without alleys). (From Pitt, R. Small Storm Flow and Particulate Washoff Contributions to Outfall Discharges. Ph.D. dissertation, Department of Civil and Environmental Engineering, the University of Wisconsin, Madison, November 1987. With permission.)


Figure 8.3 Flow sources for example medium-density residential area having clayey soils. (From Pitt, R. and J. Voorhees. Source loading and management model (SLAMM). Seminar Publication: National Conference on Urban Runoff Management: Enhancing Urban Watershed Management at the Local, County, and State Levels. March 30–April 2, 1993. Center for Environmental Research Information, U.S. Environmental Protection Agency. EPA/625/R-95/003. Cincinnati. OH. pp. 225–243. April 1995.)

Figure 8.3 is an example of a typical medium-density residential area, showing the percentage of runoff originating from different major sources, as a function of rain depth. For storms of up to about 0.1 in in depth, street surfaces contribute about one half of the total runoff to the outfall. This contribution decreased to about 20% for storms greater than about 0.25 in in depth. This decrease in the significance of streets as a source of water is associated with an increase of water contributions from landscaped areas (which make up more than 75% of the area and have clayey soils). Similarly, the significance of runoff from driveways and roofs also starts off relatively high and then decreases with increasing storm depth. Obviously, this is just an example plot, and the source contributions would vary greatly for different land uses/development conditions, rainfall patterns, and the use of different source area controls.

A major use of SLAMM is to better understand the role of different sources of pollutants and the suitability of controls that can be applied at the sources and at outfalls. As an example, to control suspended solids, street cleaning (or any other method to reduce the washoff of particulates from streets) may be very effective for the smallest storms, but would have very little benefit for storms greater than about 0.25 in in depth. However, erosion control from landscaped surfaces may be effective over a wider range of storms. The following list shows the different control programs that were investigated in this hypothetical medium-density residential area:

• Base level (as built in 1961–1980 with no additional controls) • Catchbasin cleaning • Street cleaning • Grass swales • Roof disconnections • Wet detention pond • Catchbasin and street cleaning combined • Roof disconnections and grass swales combined • All of the controls combined

This residential area, which was based upon actual Birmingham, AL, field observations for homes built between 1961 and 1980, has no controls, including no street cleaning or catchbasin cleaning. The use of catchbasin cleaning in the area, in addition to street cleaning, was evaluated. Grass swale use was also evaluated, but swales are an unlikely retrofit option and would only be appropriate for newly developing areas. However, it is possible to disconnect some of the roof drainages and divert the roof runoff away from the drainage system and onto grass surfaces for


infiltration in existing developments. In addition, wet detention ponds can be retrofitted in different areas and at outfalls. Besides those controls examined individually, catch basin and street cleaning controls combined were also evaluated, in addition to the combination of disconnecting some of the rooftops and the use of grass swales. Finally, all of the controls were also examined together.

The following list shows a general description of this hypothetical area:

• All curb and gutter drainage (in fair condition) • 70% of roofs drain to landscaped areas • 50% of driveways drain to lawns • 90% of streets are intermediate texture (remaining are rough) • No street cleaning • No catchbasins

About one half of the driveways currently drain to landscaped areas, while the other half drain directly to the pavement or the drainage system. Almost all of the streets are of intermediate texture, and about 10% are rough textured. As noted earlier, there currently is no street cleaning or catchbasin cleaning.

The level of catchbasin use that was investigated for this site included 950 ft3 of total sump volume per 100 acres (typical for this land use), with a cost of about $50 per catchbasin cleaning. Typically, catchbasins in this area could be cleaned about twice a year, for a total annual cost of about $85 per acre of the watershed. Street cleaning could also be used, with a monthly cleaning effort for about $30 per year per watershed acre. Grass swale drainage was also investigated. Assuming that swales could be used throughout the area, there could be 350 ft of swales per acre (typical for this land use), and the swales could be 3.5 ft wide. Because of the clayey soil conditions, an average infiltration rate of about 0.5 in per hour was used in this analysis, based on many different double ring infiltrometer tests of typical soil conditions. Swales cost much less than conventional curb and gutter systems, but have an increased maintenance frequency. Again, the use of grass swales is appropriate for new development, but not for retrofitting in this area.

Roof disconnections could also be utilized as a control measure by directing all roof drains to landscaped areas. The objective would be to direct all the roof drains to landscaped areas. Since 70% of the roofs already drain to the landscaped areas, only 30% would be further disconnected, at a cost of about $125 per household. The estimated total annual cost for roof disconnections would be about $10 per watershed acre. An outfall wet detention pond suitable for 100 acres of this medium-density residential area would have a wet pond surface of 0.5% of drainage area to provide about 90% suspended solids control. It would need 3 ft of dead storage and live storage equal to runoff from 1.25 in of rain. The total annual cost for wet detention ponds was estimated to be about $130 per watershed acre.

Table 8.12 summarizes the SLAMM results for runoff volume, suspended solids, filterable phosphate, and total lead for 100 acres of this medium-density residential area. The only control practices evaluated that would reduce runoff volume are the grass swales and roof disconnections. All of the other control practices evaluated do not infiltrate stormwater. Table 8.12 also shows the total annual average volumetric runoff coefficient (Rv) for these different options. The base level of control has an annual flow-weighted Rv of about 0.3, while the use of swales would reduce the Rv to about 0.1. Only a small reduction of Rv (less than 10%) would be associated with complete roof disconnections, compared to the existing situation, because of the large number of roof disconnections that already exist. The suspended solids analyses show that catchbasin cleaning alone could result in about 14% suspended solids reductions. Street cleaning would have very little benefit, while the use of grass swales would reduce the suspended solids discharges by about 60%. Grass swales would have minimal effect on the reduction of suspended solids concentrations at the outfall, but provide about 60% reductions in annual pollutant mass discharges (they are primarily an infiltration device, having very little filtering benefit). Wet detention ponds would remove about 90% of the mass and concentrations of suspended solids. Similar observations can be made for filterable phosphates and total lead.

DA

TA

INT

ER

PR

ET

AT

ION

633

lbs/acre Annual

minimal

3000

0.29

0.01 0.49

0.75 1.28

0.48

0.26

0.29

0.46

293

397

2.0 1.7

2.0

1.6

1.8

1.7

1.6

6.4

Total Lead

14

63

24 21

87 73

14

77

Flow-wtg. µg/L

543 468

14

543 0

513 6

443 18

69 87

468 14

352 35

Filterable Phosphate

lbs/acre

minimal

Annual

0.4069

0.58 0.58

0.58

0.22 0.36

0.55 0.03

0.58

0.58

0.18

N/A

N/A

333

N/A

N/A

62

25

0 0

0 0

5

0 0

0 0

Flow-wtg. µg/L

157 157

0

157 0

151 1

156 1

157 0

157 0

139 11

SLAMM Predicted Runoff and Pollutant Discharge Conditions for Examplea

lbs/acre

minimal

Annual

Suspended Solids

1430 1230

1430

1430

1250

1230

0.43

0.10

0.58

0.01

200

N/A

554 876

N/A

185

200

526 904

14

61

87

14

63

0 0

0 0

Flow–wtg. mg/L

385 331

14

385 0

380 1

410 –6

49 87

331 14

403 –5

77–100 77–100

77–100

63–100

76–100

77–100

77–100

63–100

Range CNFlow-wtg.

0.12

0.28

Runoff Volume

0.3 0.3

0.3

0.3

0.3

0.1

Rv

0.00026

Annual ft3/acre

minimal

59800

59800

59800

23300 36500

56000

59800

59800

20900 38900

3800

N/A

N/A

N/A

N/A

61

65

0 0

0 0

6 0

0 0

0 0

Birmingham 1976 rains: (112 rains, 55 in. total

reduction (lbs. or $/ft3) CB and street cleaning

0.01–3.84 in. each)

reduction (lbs or ft3)





reduction (lbs or ft3) Roof dis. and swales

Catchbasin cleaning

Roof disconnections ($minimal/acre/yr)

cost ($/lb or $/ft3)



Wet detention pond




cost ($/lb ir $/ft3)

Base (no controls)

($130/acre/yr)

($115/acre/yr)

reduction (%)

reduction (%)

reduction (%)

reduction (%)

reduction (%)

reduction (%)

reduction (%)

Street cleaning ($85/acre/yr)

($30/acre/yr)

($10/acre/yr)

(10/acre/yr)

Grass swales

Table 8.12

634 S

TO

RM

WA

TE

R E

FF

EC

TS

HA

ND

BO

OK

From Pitt, R. and J. Voorhees. Source loading and management model (SLAMM). Seminar Publication: National Conference on Urban Runoff March 30–April 2, 1993. Center forManagement: Enhancing Urban Watershed Management at the Local, County, and State Levels.

Environmental Research Information, U.S. Environmental Protection Agency. EPA/625/R-95/003. Cincinnati. OH. pp. 225–243. April 1995.

lbs/acre Annual

Medium-density residential area, developed in 1961–1980, with clayey soils (curbs and gutters); new development controls (not retrofi t).

0.05 1.98

129

Total Lead

97

Flow-wtg. µg/L

36 93

Filterable Phosphate

lbs/acre Annual

0.18 0.40

638 69

(continued)

Flow-wtg. µg/L

139 11

SLAMM Predicted Runoff and Pollutant Discharge Conditions for Examplea

lbs/acre Annual

Suspended Solids

1375

0.19

55

96

Flow–wtg. mg/L

42 89

63–100

Range CNFlow-wtg.

Runoff Volume

0.1

Rv Annual ft3/acre

0.0066

20900 38900 65

Birmingham 1976 rains: (112 rains, 55 in. total

0.01–3.84 in. each)



All above controls

($255/acre/yr)

reduction (%)

Table 8.12

a


0.007

0.006

0.005

0.004

0.003

0.002

0.001

0.000 0 10 20 30 40 50 60 70 80 90 100

Maximum percentage runoff volume reduction

roof disconnections

grass swales

roof disc. + swales

all controls

Figure 8.4 0 Cost-effectiveness data for runoff volume reduction benefits. (From Pitt, R. and J. Voorhees. Source loading and management model (SLAMM). Seminar Publication: National Conference on Urban Runoff Management: Enhancing Urban Watershed Management at the Local, County, and State Levels. March 30–April 2, 1993. Center for Environmental Research Information, U.S. Environmental Protection Agency. EPA/625/R-95/003. Cincinnati. OH. pp. 225–243. April 1995.)

0.7

0.6

0.5

0.4

0.3

0.2

0.1

0.0 0 10 20 30 40 50 60 70 80 90 100

catchbasin cleaning

CB & street cleaning

all controls

wet detention roof disc. &

grass swalesgrass swales

Maximum percentage suspended solids reduction

Figure 8.5 0 Cost-effectiveness data for suspended solids reduction benefits. (From Pitt, R. and J. Voorhees. Source loading and management model (SLAMM). Seminar Publication: National Conference on Urban Runoff Management: Enhancing Urban Watershed Management at the Local, County, and State Levels. March 30–April 2, 1993. Center for Environmental Research Information, U.S. Environmental Protection Agency. EPA/625/R-95/003. Cincinnati. OH. pp. 225–243. April 1995.)

Figures 8.4 through 8.7 show the maximum percentage reductions in runoff volume and pollutants, along with associated unit removal costs. As an example, Figure 8.4 shows that roof disconnections would have a very small potential maximum benefit for runoff volume reduction and at a very high unit cost compared to the other practices. The use of grass swales could have about a 60% reduction at minimal cost. The use of roof disconnections plus swales would slightly increase the maximum benefit to about 65%, at a small unit cost. Obviously, the use of roof disconnections alone, or all control practices combined, is very inefficient for this example. For suspended solids control, catchbasin cleaning and street cleaning would have minimal benefit at high cost, while the use of grass swales would produce a substantial benefit at very small cost. However, if additional control is necessary, the use of wet detention ponds may be necessary at a higher cost. If close to 95% reduction of suspended solids was required, then all of the controls investigated could be used together, but at substantial cost.

$/po

und

susp

ende

d so

lids

redu

ctio

n $/

cub

ic fo

ot

red

uct

ion


700

600

500

400

300

200

100

0 0 10 20 30 40 50 60 70 80 90 100

all controls

roof disconnections

roof disc. & grass swales

grass swales

Maximum percentage dissolved phosphate reduction

Figure 8.6 0 Cost-effectiveness data for dissolved phosphate reduction benefits. (From Pitt, R. and J. Voorhees. Source loading and management model (SLAMM). Seminar Publication: National Conference on Urban Runoff Management: Enhancing Urban Watershed Management at the Local, County, and State Levels. March 30–April 2, 1993. Center for Environmental Research Information, U.S. Environmental Protection Agency. EPA/625/R-95/003. Cincinnati. OH. pp. 225–243. April 1995.)

500

400

300

200

100

0 0 10 20 30 40 50 60 70 80 90 100

all controls

roof disconnections

roof disc. & grass swales

grass swales

wet detention

CB & street cleaning

catchbasin cleaning

Maximum percentage total lead reduction

Figure 8.7 0 Cost-effectiveness data for total lead reduction benefits. (From Pitt, R. and J. Voorhees. Source loading and management model (SLAMM). Seminar Publication: National Conference on Urban Runoff Management: Enhancing Urban Watershed Management at the Local, County, and State Levels. March 30–April 2, 1993. Center for Environmental Research Information, U.S. Environmental Protection Agency. EPA/625/R-95/003. Cincinnati. OH. pp. 225–243. April 1995.)

SUMMARY

This chapter has provided some, but limited, insight into how an investigator of urban receiving waters can interpret collected data and develop appropriate conclusions. This is obviously not an easy task. The main chapters of the book include many case studies and a large number of references to illustrate how prior investigators have accomplished this difficult task. The investigator should

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tota

l lea

d re

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ion

$/po

und

diss

olve

d ph

osph

ate

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n


consult selected references that are similar in location, scope, and/or objectives. Some major theses of this book are summarized below.

It is critical that the investigator have a good idea of what is to be accomplished and develop a suitable experimental design with an appropriate tiered approach. Shortcomings of many investigations can be traced to a lack of initial thought and suitable study hypotheses. In addition, while it is critical to retain flexibility and increase attention given to newly uncovered interesting phenomena, it is important not to keep changing direction based on preliminary conclusions. Of course, the reality of limited resources also precludes continued increases in project scope. Most receiving water investigations are probably only initial investigations, with little prior specific data for the location being studied, and it is natural for many new questions to develop during the studies. A tiered, weight-of-evidence approach enables the most significant objectives to be adequately addressed, with resources available for more detailed investigations to clarify issues.

It is difficult to examine the collected data and clearly identify some beneficial use impairments, especially considering the dynamic and seasonal nature of watersheds and receiving waters. It is easy to miss important short-term events and to misjudge the significance of other seemingly obvious events. Careful and complete sampling, especially if conducted using a stratified random sampling strategy, can help reduce these problems. Well-calibrated and verified models are also important because they allow a long-term perspective of the discretely collected data.

Cause-and-effect relationships tying together stressors and biological impairments are especially difficult to identify and quantify, requiring specialized tests and the weight-of-evidence approach. The use of adequate and appropriate reference sites is very important for biological evaluations, as comparisons to water quality criteria are uncertain for stormwater-related problems, habitat guidance is in its infancy, and contaminated stream sediment guidelines are unclear. The use of available criteria is needed, obviously, but criteria exceedances should be considered red flags to focus site-specific investigations. Human risks should be much more closely related to water quality criteria for water contact, drinking water supplies, and fish consumption, but there is still a potential for error when predicting actual exposure associated with stormwater sources.

The implications of receiving water investigations can be extremely important, especially if remedial action is warranted or if problems are not to be worsened in the future. It is therefore necessary that the whole watershed and associated urban infrastructure be considered. It would be very unusual to find an urban or agricultural receiving water in a completely developed watershed that has significant acceptable uses in the absence of dramatic stormwater controls. In most cases, these streams are managed, it at all, solely for flood control and drainage, with no acknowledgment of other reasonably acceptable uses of noncontact recreation and biological life. Unfortunately, many urban and agricultural streams attract children who play around and in them. Obviously, water-contact recreation and fish consumption are not appropriate uses of most urban and agricultural streams, but these uses do occur, often by members of the community most at risk. These concerns and challenges can be effectively addressed by linking assessments of stormwater effects with progressive watershed management.

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