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Risk Assessment of Radon in Drinking Water (1999)

Chapter: 8 Mitigation

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Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
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8
Mitigation

Mitigation Of Radon In Indoor Air

Radon Entry into Buildings: A Brief Review

Radon is a ubiquitous constituent of soil gas as its radioactive parent, 226Ra, is widely distributed in the earth's crust. Typical soil-gas radon concentrations are around 30,000-300,000 Bq m-3, and values ranging from about 5,000 Bq m-3 to about 5,000,000 Bq m-3 have been reported. The principal mechanisms of radon transport in porous media (e.g., soil) are advection and diffusion; both are sources of radon entry into buildings, and they are described briefly in this section. More complete discussions of radon transport in soils and entry into buildings can be found in the literature (Sextro 1994; Nazaroff 1992; Nazaroff and others 1988).

Advection

Bulk flow of soil gas that contains radon is the main mechanism of radon entry into buildings. This flow occurs in response to pressure differences between the air in buildings and the air in the adjacent soil. These differences are established by the natural interaction between the building and the surrounding environment and in some cases by the operation of mechanical systems within the building.

The temperature difference between the air in a building and the air outside creates a pressure gradient across the building shell that varies with height along the shell. When the indoor air temperature is higher than the outdoor, the indoor air pressure in the lower parts of the building (e.g., in the basement or the region of the ground-contact floor) is slightly lower than the air pressure in the adjacent

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
×

soil; conversely, at the top of the building, the pressure gradient across the building shell is reversed, so air flows out of the building. This thermal-stack effect is one of two principal mechanisms responsible for the natural ventilation of buildings; it is sometimes referred to as infiltration.

Wind also creates pressure differences between the inside and the outside of a building. The pressure fields can be complex and depend on the size and shape of the building and the wind direction. The pressure fields also extend into the surrounding soil, increasing the pressure of the soil air on the upwind side and decreasing it on the downwind side of the building (Riley and others 1996). The net effect is usually an airflow out of the top of the building caused by the Bernoulli effect of the wind over the roof or by the reduction in air pressure on the leeward side of the building. In response to the slightly lower air pressure in the building, ''makeup" air flows in through openings in the building shell, some of which might provide direct contact with soil air.

The effects of the thermal stack and the wind can independently result in indoor-outdoor pressure differences of about the same size at the lower portions of the building shell. For example, an indoor-outdoor temperature difference of 20 °C (a common wintertime temperature difference in many parts of the United States) results in an indoor-outdoor pressure difference of about-3 Pa at the bottom of the thermal stack (the basement or other ground-contact floor). Similarly, a wind speed of 4 m s-1 results in an indoor-outdoor pressure difference of about-2 Pa for a typical house (the relationship between the indoor-outdoor pressure difference and wind speed is quadratic, so doubling the wind speed increases the pressure difference by a factor of 4). Those are "steady-state" values. Temperature differences usually do not change very rapidly, but in the course of a day outdoor temperatures can change by 20 °C or more as part of the diurnal cycle. Wind speeds and directions are highly changeable, and this leads to substantial variation in "instantaneous" pressures. A more detailed discussion of the pressure gradients developed in buildings can be found in Liddament (1986).

The operation of mechanical systems in a building can lower the pressure in a building, especially when the flows induced by these systems are unbalanced. Operation of an exhaust fan—such as a bathroom or kitchen fan, whole-house fan, or, in some cases, an attic fan—will result in lower indoor pressures. Just as in the case of infiltration, the "makeup" air flows into the house through leaks in the building shell, some of which provide a pathway for soil-air entry. Operation of a forced-air heating and cooling system can also lead to unbalanced flows and result in lower indoor pressures, depending on the locations of the supply and return ducts and their leakage characteristics.

Diffusion

Molecular diffusion, driven by the concentration difference between low-concentration regions, such as the interior of a building, and the higher-concen-

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
×

tration soil is another mechanism for radon entry into buildings. A key controlling variable—in addition to the radon concentration gradient—is the diffusivity of any material that separates the soil from the building interior, such as a concrete floor. In the case of an open soil floor, and in the absence of pressure differences, the radon flux density is the highest across this interface and is about the same as would be observed outdoors (soil moisture differences can have an effect on the diffusivity of soil). For soils with typical radium content, the radon flux density is 1–2 × 10-2 Bq m-2 s-1.

The presence of a concrete floor can increase the concentration gradient over that found in open soil, but radon diffusivities are typically smaller in concrete than for soil. The concrete floor acts as a diffusion barrier; diffusive radon entry though such a floor is likely to be somewhat lower than that for open soil. For nominal values of the diffusivity of concrete and typical radon concentrations in soil gas adjacent to a building, the radon entry rate due to diffusion through a concrete floor is about 1 × 10-2 Bq m-2 s-1, which is about half the open-soil value. Most of this radon is from the soil itself, as opposed to the radon arising from radium in the concrete (Sextro 1994). This estimate is consistent with measurements of flux density conducted as part of extensive field experiments, where the average flux density was 1.3 × 10-2 Bq m-2 s-1 (Turk and others 1990).

Building materials themselves—especially those with soil-based constituents, such as concrete, brick, and natural stone—contain radium and will thus be a source of radon diffusion into indoor air. In most cases, however, the amount of radium in such materials is small enough that, in combination with the diffusivity of the material and typical infiltration and ventilation rates of buildings, their overall contribution to indoor radon concentrations is modest.

Other Sources

Three other sources of radon are worth noting. The first is advective transport of soil gas driven by changes in atmospheric pressure. Although large changes in atmospheric pressure can result from changes in weather, they are relatively infrequent compared with the smaller diurnal and semidiurnal atmospheric pressure changes (Robinson and Sextro 1997). Overall, these effects are estimated to be small and to yield overall radon entry rates roughly the same as that due to the second source, infiltrating outdoor air. The latter, considered in more detail in chapter 2, provides an irreducible "baseline" indoor radon concentration.

The third source is the topic of this report: indoor use of water that contains dissolved radon, which is the subject of detailed discussions elsewhere. In the context of other sources, the average contribution made by water to indoor-air radon concentrations is very modest, given that the average transfer coefficient is 10-4.

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
×
Radon Entry in Context

Several sources of indoor radon have been described; from all but outdoor air, the resulting indoor concentrations (and hence exposures) depend on the combination of the source strength and the ventilation rate of the building. As noted earlier, the stack and wind effects are primarily responsible for the natural ventilation of buildings, in addition to providing a driving force for radon transport into buildings. It is useful to provide a context for these flows.

The steady-state solution to the first-order differential equation that describes indoor radon concentrations illustrates the key variables:

Where,

Ci is the indoor concentration (Bq m-3), Co is the outdoor concentration (Bq m-3), S is the radon entry or production rate (Bq per unit time, t), V is the house volume (m3), R is the removal rate (t-1).

Here R can account for any method of removal and is just the sum of the individual removal terms. In this case, R is the air-exchange rate (AER). A typical single-story house has a "footprint" of 120 m2 and a corresponding volume of about 300 m3. Annual average natural ventilation rates are about 0.9 h-1. Using those values for V and AER, the air flow rate (the product of V and AER) through this house is 270 m3 h-1.

The radon entry rate corresponding to the Environmental Protection Agency (EPA) guideline concentration of 150 Bq m-3 can be estimated from equation 8.1. Neglecting any contribution from outdoor air and using the ventilation rate described above, the radon entry rate, S, is about 40,000 Bq h-1 (about 11 Bq s-1). That can be compared with the estimates of diffusive radon entry. Assuming a floor area of 120 m2, the entry rate due to diffusion is about 1 Bq s-1, a small fraction of what is needed to produce an indoor air concentration of 150 Bq m-3.

Similarly, the soil-gas flow to produce this indoor concentration can be estimated. Assuming a typical value of about 40,000 Bq m-3 for the concentration of radon in soil gas, the soil-gas entry rate is about 1 m3 h-1, which is about 0.4% of the overall air flow rate into the house.

Mitigation Methods for Existing Houses

Conceptually, there are two approaches for mitigating indoor radon concentrations (or most other indoor pollutants, for that matter): source control and

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
×

concentration reduction. One can use equation 8.1 to provide some insight into the relative efficacy of the two approaches. Indoor radon concentration is directly proportional to the source term and (again neglecting outdoor air as a source) inversely proportional to the removal terms. Considering the latter first, removal can mean either increased ventilation or some other method of removing radon or radon decay products from indoor air. In any case, for the previous example, to decrease the radon concentration by a factor of 2 by ventilation alone, the AER will need to be increased to 1.8 h-1. Although that is not an excessive ventilation rate and is often achieved naturally when doors and windows are open, AER values of 2 h-1 commonly have comfort and energy penalties during colder seasons. Thus, this means of reducing radon concentration has some practical upper limits. In addition, forced ventilation can result in additional depressurization of a building and potentially increase the radon entry rate.

Other nonventilation removal methods are possible, and two are described in more detail below. As with ventilation, substantial removal means processing indoor air at rates that are comparable with or greater than the ventilation rate (about 270 m3 h-1 in the example above). It also means that essentially the entire living space will need to be treated; this could require multiple single-room reduction devices (such as air cleaning, described below) or whole-house devices used in conjunction with a forced-air system. In the following sections, source-control methods are described first and then concentration-reduction methods.

Source Control

When high indoor radon concentrations in houses were found in various locations in North America in the middle 1970s, initial research on reduction methods was based on two key assumptions: that the source term was high concentrations of radium in soil materials derived from uranium mining and that the principal means of radon transport and entry was diffusion. Thus, initial attempts at source control focused on removal of the materials, typically uranium mill tailings used as back fill under floor slabs or adjacent to basement walls. In addition, several projects investigated the use of coatings and other sealants that would serve as an additional barrier to radon diffusion (see, for example, Culot and others 1978). Although removal of some of the high-radium-concentration materials had an effect, indoor radon concentrations in some cases were not reduced commensurately. As additional measurements of indoor radon concentration were conducted, houses were found with high indoor radon concentrations that had no known anthropogenically enhanced radon source (Sachs and others 1982). At the same time, mass-balance considerations (similar to equation 8.1 above) showed that diffusion alone had only a slight potential to produce the high indoor concentrations that were being observed (Bruno 1983).

Although removal of high-radium-concentration source materials can be part of an overall radon-control method, it generally is not part of current practice,

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
×

because it is not necessary (except in rare cases). The vast majority of radon-mitigation systems now installed in existing houses rely on mechanically driven, or active, subslab depressurization (ASD) techniques (Henschel 1994). These methods seek to reverse the pressure gradient across the part of the building shell that is in contact with the soil. As noted earlier, this pressure difference drives the advective flow of radon-bearing soil gas into a building. The systems are sometimes referred to as subslab ventilation systems, but as a general rule that is a misnomer. When operated in a depressurized mode, the system does draw some outdoor air from the surface into the soil near the building. It also draws air into this region from the basement (reversing the flow of gas in the cracks and openings in the building shell). The flow of air may dilute the soil-gas radon concentration in the vicinity of the building somewhat, but the extent depends on the permeability of the soil. The key operating principle is still reversal of the indoor-outdoor pressure gradient.

Operationally, a subslab system consists of one or more pipes that penetrate the floor slab. The pipes, typically 7–15 cm in diameter, run vertically through the house and terminate above the roof. A mechanical fan, usually an in-line axial fan designed specifically for this application, is installed in the pipe system where it passes through the attic or some other location outside the conditioned living space of the house. The fan operates at about 100–400 m3 h-1 at a pressure of up to a few hundred pascals (Henschel 1993).

In an ASD system, the fan creates a low-pressure zone in the soil outside the building shell. A successful system will reverse—or at least reduce—the pressure gradient at all major building-shell penetrations that are in contact with the soil. An important entry pathway for soil gas in many basement structures is the expansion-contraction joint at the edge of the concrete floor slab where it abuts the wall. In some cases, there will also be openings or utility penetrations through the basement walls or, as in the case of walls constructed of hollow-core "cinder" or concrete block, the wall itself is permeable to air flow. To eliminate or reduce soil gas entry in these areas, the low-pressure zone must extend beyond the region of the floor and up the walls. Almost all the retrofitted ASD systems are successful in reducing indoor radon concentrations to less than 150 Bq m-3 and often concentrations are reduced to about 75 Bq m-3. In some cases when the basement walls are constructed of blocks, depressurization pipes are inserted into the hollow cores of the blocks themselves. Because these cores are typically interconnected, directly or through thin permeable concrete "webs," there is in effect a depressurized plenum within the walls themselves, thus largely eliminating any flow of soil gas across the wall and into the building interior.

Over the last decade, a considerable amount of research and practical experience on the installation of these systems has been accumulated (Henschel 1993). Two elements aid the successful implementation of an ASD system. One is the presence of a high-permeability gravel layer below the floor slab. This layer essentially establishes a low-flow-resistance pressure plenum that enhances the

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
×

lateral extension of the pressure field, as described earlier. In some cases, the presence or continuity of the gravel layer cannot be easily determined. In these cases, a ~ 1 m diameter sump or pit is dug into the soil below the slab at the point where the ASD pipe extends below the slab. The pit helps to ensure that the pressure field created by the ASD extends as far as possible throughout the region of the soil-building interface.

The second practical element in the implementation of an ASD system is sealing as many of the potential radon-entry locations as possible. Although sealing by itself is usually not effective in eliminating radon entry, sealing does enhance the effectiveness of an ASD system because it helps to reduce any short-circuiting of air flow from the building interior into the depressurized region below the floor slab. By reducing this air leakage, the low-pressure field created by the ASD system can be further extended laterally along the soil-building interface.

One variant of the subslab system uses a fan to pressurize the region below the floor slab. In this case, the system is providing ventilation of soil gas, thus reducing the radon concentration in the soil region adjacent to the building. Rather than reducing or reversing the pressure gradient across the building shell, this method actually increases the interior-to-exterior pressure difference and so increases the flow of gas from the soil into the building. When successfully implemented, the reduction in radon concentration in the soil gas more than compensates for the increased flow. Careful studies have shown, however, that high soil permeability is key to the successful use of this technique, because it permits a larger dilution effect (Gadgil and others 1994; Turk and others 1991a; 1991b).

Basement pressurization has also been used to control radon entry. This method uses the same principle for control as does an ASD system, but it pressurizes the entire basement volume to reverse the indoor-outdoor pressure gradient. Successful use of the technique in a research-house study provided strong empirical evidence that radon entry into buildings is dominated by advective transport. However, as a practical matter, use of the technique has been limited to basements that can be made very tight with respect to air leakage, particularly the membrane between the basement and first floor. Pressurization is done with conditioned air, usually drawn from the first floor. If flow rates are too large, a substantial energy (and in some cases comfort) penalty is associated with heating or cooling the extra "make-up" air as it infiltrates into the house. This method can also create backdrafting problems for fireplaces or other combustion appliances on the first floor (Turk and others 1991a; 1991b).

The source-control methods described thus far for use in existing houses are all mechanically driven (that is, fan-powered), so-called "active" methods. Two other techniques—both passive—have been used. The first technique, sealing, has been noted earlier. Empirically, this method has not been found to reliably produce substantial reductions in radon entry, largely because it is often difficult

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
×

to find or satisfactorily seal all the leakage pathways. The problem is acute when part of the soil-structure interface has a low resistance to flow, as when there is a gravel layer below the floor or when the basement walls are constructed of hollow-core block. The second passive (nonmechanical) technique that has been used as a retrofit mitigation system with some (but not uniform) success is the passive thermal stack. Similar in some respects to ASD systems, it consists of a pipe system that is inserted through the floor, and passes through the house and out through the roof. It is important that the pipe pass through the heated portion of the house, because it relies on heat transfer from this conditioned space to heat the air column inside the stack, thus creating the thermal stack effect. There is a small amount of pressure loss at each bend in the pipe, so it is also important to minimize the number of bends in the pipe system as it passes through the house.

For this system to be effective, the pressure field developed by the stack below the floor slab needs to be sufficient to reverse or at least substantially reduce, the pressure gradient between the soil and the building interior, which drives advective flow of gas from the soil into the building. The soil-to-building-interior pressure difference will be greatest when the inside-the-stack-to-outdoor temperature difference is the largest, for example, during the winter in cold or moderate climates. This is the same period when the advective transport of soil gas into the building is potentially the greatest. The influence of wind can complicate the behavior of a passive stack system. As described earlier, wind can depressurize the building interior, in addition to the depressurization caused by the stack effect. Wind can also affect the flows and pressures at the stack opening, depending on the wind direction with respect to the orientation of the roof.

It is important to have a high-permeability zone below the floor to ensure that the pressure field created by the passive stack extends along the soil-building interface, especially inasmuch as the pressure field generated by the stack is typically 1–10 Pa less than the air pressure inside the building, compared with the 100-to 400-Pa pressure difference generated by an ASD system (Gilroy and Kaschak 1990). In an existing house, the presence of such a layer and the extent to which it is present throughout can be difficult to determine.

Concentration Reduction

Unlike source-control methods, which seek to limit radon entry, concentration-control methods are designed to reduce radon or radon decay-product concentrations in indoor air. Three concentration-control techniques will be described in this section.

As mentioned earlier, increased ventilation can reduce both radon and radon decay-product concentrations, as long as it does not enhance the indoor-outdoor pressure difference. In one set of experiments conducted in a house, basement radon concentrations were observed to be lower when the basement windows were open. Measurements conducted with a tracer gas showed that basement

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
×

ventilation increased somewhat with the windows open but that the largest effect was due to reducing the indoor-outdoor pressure difference across the basement wall (Cavallo and others 1996).

One method of increasing ventilation while avoiding some of the energy and comfort penalties noted earlier is to use an air-to-air heat-exchanger system, often referred to as a heat-recovery ventilation (HRV) system. In this approach, designed around commercially available HRV units, ventilation air is exhausted through a heat exchanger through which incoming unheated air also passes. The heat-exchange process is about 40–80% efficient thermally and substantially reduces the energy cost of increased ventilation (Turk and others 1991a; Fisk and Turiel 1983). Such systems have been used successfully for radon control, especially in houses with basements. Because most radon entry occurs through the basement floor and walls, basement radon concentrations are often higher than elsewhere in a house. Use of an HRV to reduce radon concentration in this space, as opposed to the whole house, means that the effective ventilation rate of the space is higher (which affords more control); by controlling basement concentrations, it also reduces radon levels throughout the house. One important ancillary benefit is that the HRV can be used to alter the basement pressure somewhat and will thus provide some additional radon-concentration reduction via source control (Turk and others 1991a).

Another method that has had very limited use is based on the sorption of radon gas by activated carbon (Bocanegra and Hopke 1989; Brisk and Turk 1984). A commercially available device based on this approach consisted of two carbon beds; one removed radon from indoor air flowing through it while the other was being purged of accumulated radon by having outdoor air passed through it and exhausted to the outdoors. The two beds were switched periodically so that the freshly purged bed was used to accumulate radon and the bed used for sorption began to be purged (Wasiolek and others 1993). Overall performance of this method is limited by the rate of air flow through the device, which in turn helps to determine the charcoal bed thickness. Like the HRV system, this approach appears to have the greatest applicability in radon control for a basement.

The third concentration-reduction approach is the use of air-cleaning to reduce radon decay-product concentrations. Unlike their chemically inert parent, the decay products 218Po, 214Pb, and 214Bi are metals and easily attach to the surfaces of any aerosols that are present (the ''attached" mode). Some decay-product atoms, particularly 218 Po, can also remain as ultrafine aerosols (the "unattached" mode, a few nanometers in diameter). Indoor air concentrations of both modes can be reduced by using an air cleaner designed to remove particles.

There have been a number of evaluations of air-cleaning systems undertaken in test chambers or actual indoor environments (reviewed in Hopke and others 1990). Some of these systems can effectively remove radon decay products from indoor air. However, the reduction of 218Po is not as large as that of 214Pb and

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
×

214Bi. At the same time, the particles are removed from the air. As a result, the unattached fraction of airborne activity increases, especially of 218Po. Because the unattached fractions of the radon progeny have been considered to be far more effective in depositing their radiation dose to lung tissue, concerns have been raised regarding the efficacy of air-cleaning as a means of mitigating the hazards arising from indoor radon.

A major problem in the previous studies was that the systems used to measure radon progeny were not able to determine the full size distribution, especially in the size range below 10 nm. Estimates of the unattached fractions were made with systems that provide a poorly defined size segregation (Ramamurthi and Hopke 1989). In many cases, the size-measurement methods and results were not clearly stated.

In 1990–1992, a research program supported initially by the New Jersey Department of Environmental Protection and then also by EPA undertook field studies to investigate the effects of room-air cleaners on radon progeny concentrations and activity-weighted size distributions (Hopke and others 1993; Wasiolek and others 1993; Li and Hopke 1992; 1991b) A unique, semicontinuous graded screen-array sampling system (Ramamurthi and Hopke 1991) was used to measure the radioactivity associated with indoor aerosol particles in the size range of 0.5–500 nm.

In an early set of studies, particles were produced by a variety of activities, such as cooking, smoldering of a cigarette, burning a candle, and operating a vacuum cleaner. Aerosol behavior in the absence of an air cleaner was determined for each condition (Li and Hopke 1991a). The experiments were then repeated with a high-efficiency filter system operating (Li and Hopke 1992). It was found that the filtration unit reduced the airborne activity concentrations by removing particles, but the reductions in estimated dose were much smaller than the decrease in PAEC.

Other experiments in normally occupied houses have involved the measurement of the effectiveness of the filtration unit and an electrostatic precipitator by comparing the cumulative frequency distributions of measurements made during a week while a particular cleaner was operating and measurements made during a background week in which no cleaner was being used (Li and Hopke 1991b). A similar experimental design was used to study the two cleaners and an ionization system in an occupied home (Hopke and others 1993). The results of the 1992 measurements in Parishville, NY, in which two ionizing units were measured along with two filtration units were described by (Hopke and others 1994). More detailed studies of the NO-RAD ionizer system under the controlled conditions of a room-sized chamber at the Lawrence Berkeley National Laboratory were performed, and there are several other ionizer-based cleaners for which there have not yet been field studies (Hopke 1997; Hopke and others 1995b).

From the more recent studies on air cleaners and their effects on exposure to and dose from airborne radon decay products, several important conclusions can

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
×

be drawn. With the new dosimetric models that more accurately reflect nasal and oral deposition of ultrafine particles, it is extremely unlikely that an air cleaner can reduce exposure and increase dose as suggested by Maher and others (1987), Sextro and others (1986), and Rudnick and others (1983). Thus, there is no reasonable likelihood that the use of an air cleaner will increase the hazards posed by indoor radon.

In studies of different types of air-cleaning devices, reductions in exposure have always exceeded reductions in dose. However, cases have been observed in which there has effectively been no reduction in dose. Thus, for many air cleaners, the clean-air delivery rate is insufficient to provide substantial protection from the radon decay-product hazard. The air cleaner might be effective in removing other contaminants—including cigarette smoke, dust, pollen, and spores—and thus provide a considerable benefit to an occupant without lowering the radon-progeny risk substantially and, more important, without raising that risk at all.

The experiments with the newest systems suggest that the combination of substantial air movement and ionization could provide sufficient reduction in exposure and dose to be effective in reducing the radon-progeny risk at radon concentrations up to around 400 Bq m-3. If it is desirable to reduce the risk to that equivalent to the average dwelling in the United States, then such units would be useful only for lower 222Rn concentrations. There would also need to be multiple units in a home to provide complete room-to-room reduction.

Mitigation Methods for New Construction

Radon-mitigation methods for new buildings can be incorporated directly into the construction process and both enhance the performance of the system and reduce the cost of installation, compared with the cost of retrofit mitigation methods. Systems for controlling radon concentrations described earlier have essentially the same applicability whether their use is in existing or new buildings, and they will not be discussed further in this section. Some cost savings might be associated with installation of systems like an HRV during the construction process or with integrating such a system into the space-conditioning system of the building.

In the following sections, the application of "existing-house" techniques for radon-entry control is discussed briefly and then a more systemic approach for making buildings radon-resistant as part of the construction process is discussed.

Application of Existing-House Radon-Entry Control Methods

As described earlier, one of the most widely used radon-mitigation techniques is ASD. Key to the successful implementation of these systems is reversal of the pressure gradient at all the major soil-gas entry points. Typically, this

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
×

requires a high-permeability zone (such as gravel) below the floor and, in some cases, around the foundation footings so that the pressure field extends up the basement walls. Sealing major openings, such as at the joint of the basement floor and the wall usually is also necessary to ensure that there are few flow ''short-circuits" that will degrade the pressure field. Achieving these in existing buildings can sometimes be problematic because the extent of the high permeability zone might not be known, although flow and pressure can be measured to provide a coarse assessment. In addition, it can be difficult to identify all the major soil-gas flow pathways through the building shell.

In new construction, both those problems are more readily addressed. The extent and quality of a gravel bed, for example, can be specified as part of the building design and as part of the construction-inspection process. Many of the leakage paths can be eliminated through design (for example, minimizing utility penetrations of ground-contact floors or walls), materials use (for example, use of low-shrinkage concrete for floors), and construction practices (for example, adequate sealing of utility penetrations).

One of the important benefits of these methods is that passive-stack methods might become more applicable. In some cases where wind or other effects might increase the depressurization of a house, thus potentially overriding the reverse pressure gradient established by the passive stack, the use of low-power fans for mechanical stack depressurization is attractive. Such systems have been tested in a limited number of homes and show promise (Fisk and others 1995; Saum 1991).

Radon-Resistant Buildings

Most of the elements required for making a building radon-resistant have already been described. In principle, if all entry routes through which soil gas can flow are eliminated or the pressure gradient that drives air flow through such openings is reversed, advective transport will not contribute to indoor radon concentrations. If successfully implemented, this approach can be achieved without the use of mechanical systems—it will constitute a so-called passive radon-resistance system. Such an approach has two important advantages over active radon-control systems: there are no mechanical or electric components to fail (for which the building occupants must maintain an awareness), and there is no concomitant energy use. On the other hand, the operation of mechanical systems can be easily monitored, for example, with a pressure gauge. Failure of a fan would, in principle, be easily detected by a change in pressure in the radon-mitigation pipe. The potential system-failure modes in a passive system are likely to be more subtle, such as those induced by the differential settling, cracking, and aging of building components, particularly foundation walls, footings, and floors.

Radon-resistant features, including those designed to reduce or eliminate radon-transport pathways and those in some cases, designed to reverse or decrease the differential driving pressure, have been proposed or incorporated into

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
×

codes and guidelines for new buildings. In addition, some builders in various parts of the United States have voluntarily adopted construction practices that they believe will limit radon entry into new homes (Spears and Nowak 1988). In 1991, Washington state adopted a radon provision as part of the Washington State Ventilation and Indoor Air Quality Code (WSBCC 1991). It provides specific details for houses built with crawlspaces for the entire state and specifies radon-resistance features for eight counties thought to have potential for high indoor radon concentrations. The specifics include use of aggregate and a membrane below the floor slab, sealing of all floor penetrations and joints, and the use of a passive stack extending from the subslab through the heated portion of the house and exhausting through the roof. At the time of its inception, the code also required provision of a long-term radon monitor in each new home; this part of the code is no longer in force.

In March 1994, EPA published a set of standards and techniques for construction of radon-resistant residential buildings (EPA 1994a). These provisions, along with EPA's county-by-county radon zone map of the United States, were incorporated as a recommendation in the Council of American Building Officials residential building code and have been adopted, with modifications in some cases, by various local building-code authorities around the country.

In 1989, Florida initiated the Florida Radon Research Program (FRRP), which was designed to be a comprehensive program of research to examine many of the details involved in typical residential construction practice and how they might be modified to provide resistance to radon entry (Sanchez and others 1990). Many of the features are directed toward reducing radon entry in the elevated slab-on-grade construction method used widely in that state. Particular attention has been paid to attempting to ensure the integrity of the floor slab and to provide a subslab membrane that is sealed at all floor penetrations (for example, plumbing pipes) and at the edge of the slab.

The FRRP, conducted in cooperation with EPA, has produced the most extensive research on radon-resistant new construction to date. Most of the features in the proposed code have been evaluated, but in only a small number of houses (see, for example Fowler and others 1994; Hintenlang and others 1994; Najafi and others 1993; Najafi and others 1995). One part of the FRRP was the development of a radon-potential map of the state, delineating regions where no special radon controls are needed, regions where only passive (radon-resistance) features are required, and regions where both radon resistance and ASD are needed (Rogers and Nielson 1994).

Results of several house-evaluation studies conducted as part of the FRRP have shown that most of the houses built in conformity with the proposed standard appear to have short-term indoor radon concentrations below 150 Bq m-3. However, there are a number of important limitations regarding these results, the most important of which might be the timing and duration of the indoor radon testing. With few exceptions, the indoor radon was measured after construction

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
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was completed but before occupancy—usually within a period of a week or two. As a result, house ventilation rates might not be similar to those during occupancy, nor will the short-term measurements of radon concentration be representative of a longer-term average indoor radon concentration.

A more generic evaluation of the effects of the different elements of the passive mitigation system proposed by the FRRP was conducted through the use of a radon-transport model with some "calibration" against data obtained in several house-evaluation projects (Nielson and others 1994). The model could not, of course, simulate failures, only the presence or absence of specified resistive features, openings through the floor, and so on. However, the results do provide insight into the relative importance of some features as applied to typical Florida residential construction. Most important was use of a vapor barrier below the floor slab, including particular attention to the details of treatment at all slab penetrations and at the slab edge, avoidance of floating slab construction, limiting concrete slump, and sealing all slab penetrations, openings, and large cracks. Interestingly, in this analysis, the presence of a passive stack to depressurize the subslab region was rated less effective than the other features of the radon-resistance system (Nielson and others 1994; Rogers and Nielson 1994). That is due in part to the assumed effectiveness of the features thought to block radon entry by reducing or eliminating air pathways between the subslab region and the house interior. In addition, the differential pressures generated across the slab by passive stacks in the houses are limited by the relatively small driving forces established by the passive stack. These occur for two main reasons: there is a narrow range of temperature differences between indoors and outdoors for most of the year, and many of the buildings are single-story (with no basement). Both limit the influence of the thermal-stack effect.

Data on the effectiveness of radon-resistance systems in other parts of the country are very sparse, and decidedly mixed. In several studies, totaling about 80 houses, measurements were made in radon-resistant houses with the passive stack closed and then after the stack was open (uncapped). Such studies appear to be based on the premise that radon-entry rates when the passive stack is closed would be similar to those observed in houses not built with radon-resistance features. However, because the passive stack is only one element in the radon-resistance system, negating its effect with a cap should not substantially affect the behavior of other parts of the radon-resistance system. More importantly, it is not clear how the radon-entry potential is affected by the use of high-permeability materials.

In one study, 46 homes were investigated in eight states (Dewey and others 1994; NAHB 1994); 41 in counties designated by EPA as having high radon potential. However, soil-gas concentrations were measured at 38 homesites (33 of them in counties designated as having high radon potential), and only 16 had soil-gas concentrations above 37,000 Bq m-3 (in one case the measurement was made at an adjacent site). Several researchers have suggested that soil-gas radon

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
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concentrations of at least 37,000 Bq m-3 (which is close to the US average soil-gas radon concentration) are needed to provide an adequate test of radon resistance in houses. This concentration was used as the selection criterion for many of the houses evaluated as part of the FRRP because it was felt that lower soil-gas radon concentrations would not test the system (Fowler and others 1994).

Of the 16 homes, four had basement radon concentrations greater than 150 Bq m-3, and one of these also had radon concentrations measured on the first floor greater than 150 Bq m-3 (all these values were measured with the passive system fully operational, that is, with the passive stack open). In each case, these were single sets of measurements, conducted with the passive stack first closed, then open. In four of the 16 cases, the stack-open measurements were conducted in the summertime, when driving forces for both radon entry and passive-stack operation are minimal.

Other studies of radon-resistant construction have had similar results and limitations (Saum 1991; Brennan and others 1990; Saum and Osborne 1990). None of these studies had a non-radon-resistant baseline against which to evaluate the overall effect of the techniques. In many cases, soil-gas concentrations were not measured, so it is difficult to determine the radon-entry potential for the specific houses. Most of the studies conducted stack-open versus stack-closed evaluations.

Overall, the inclusion of the passive stack had the largest effect in houses with basements. In some cases, the short-term measurements suggest reductions as great as 90%, compared with stack-closed basement radon concentrations. On the average, reductions of about 40% are more typical. For slab-on-grade houses, the effect of the passive stack is considerably reduced. Comparing the stack-open and stack-closed measurements from the few houses for which there are data reveals no discernible effect. In part, that is due to the low soil-gas radon concentrations in one study. Thus, with one exception, all the initial stack-closed indoor radon concentrations are low. In the FRRP studies, soil-gas radon concentrations were often high (this was one criterion in the choice of study houses), but, as noted earlier, the passive stack was thought to be a less integral part of the radon-resistance system.

Three key questions remain unanswered:

  • How well do houses built to be radon-resistant perform when compared with those built without radon-resistance features?

With one exception, no side-by-side comparisons address this question. A study of 89 homes in two areas near Colorado Springs was conducted in which 54 new homes were built within two subdivisions; 35 homes built without any radon-resistance features served as "control" homes. Of the 54 homes, 12 were tested as radon-resistant (the remaining 42 used active, mechanical systems for radon control). Radon was measured in the basements with 2-d open-face charcoal canisters, all on the same 2 days in December to remove weather effects

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
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(Burkhart and Kladder 1991). In the original paper, the average indoor radon concentration of the 12 homes was not statistically different from that of the control homes, as analyzed either by area or in combination. Further analysis of 11 of the 12 homes examined the radon-resistance features of each home. Only five had passive stacks in addition to sealing and other techniques used to limit entry of radon; of these, only three had stacks passing through the heated portion of the homes. The three passive-stack homes averaged 130 Bq m-3 over the 2 days when radon was measured; one home had slightly over 150 Bq m-3. The control homes averaged 450 Bq m-3 over the same period (Kladder and others 1991).

The only other study that has attempted to estimate the average indoor radon concentration with and without radon-resistance construction features was done as part of EPA's new-house evaluation program (Murane 1998). Measurements were made in 148 houses that were built with various radon-resistance features. All but five of the houses were built near Denver and Colorado Springs, CO and did not include a passive stack as part of radon control. The other five, built near Detroit, MI, did include a passive stack. Radon measurements were made in these houses—usually in the basement and in some cases on the first floor as well—with open-face charcoal canisters with a 2-d sampling period. The results were tabulated by ZIP code and compared with measurements in the same ZIP code done as part of the Colorado and Michigan state radon surveys (also done with short-integration time-measurement techniques). The state surveys were done in occupied houses during the winter months, but some of the new-house program measurements were done in other seasons. It is not clear whether the new houses were occupied at the time the indoor radon was measured. Overall, in the 143 Colorado houses, the average basement radon concentration was 190 Bq m-3, compared with 230 Bq m-3 in 94 control houses in the same ZIP codes. In the case of the Detroit-area houses, there were only five radon-resistant houses and four controls; basement concentrations averaged 90 and 50 Bq m-3, respectively.

Comparing the two sets of houses within each ZIP code produces considerable variation in whether the average radon concentration in the radon-resistant houses is lower or higher than of the average measured in the "control" houses. Given that all the results are based on short-term measurements, with some seasonal differences in measurements between the "control" houses and the radon-resistant ones, comparison of the average radon concentrations in the two sets of houses does not provide a strong basis for evaluating the effectiveness of the radon-resistance features. Furthermore, because one is looking for small differences between the average concentrations in the two sets—perhaps a factor of 2 or 3 at most—the study and the measurement techniques will have to be carefully designed and executed to ensure statistically meaningful results.

  • Can radon-resistance systems be relied on for the lifetime of a building?

As has been indicated in the earlier discussion, there are no data on the long-

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
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term behavior of houses built with radon-resistance features. One house in Florida built and tested as part of the FRRP was revisited as part of an examination of the durability of active radon-mitigation systems. At the time of construction, provision was made for installation of an active system because the soil-gas concentrations were about 60,000 Bq m-3. The postconstruction indoor radon concentration was 60 Bq m-3, so an active system was never installed. The revisit, only 16 mo after construction, found essentially the same indoor radon concentration (70 Bq m-3), on the basis of the results of two long-term alpha-track measurements over periods of 4 and 5 mo (Dehmel and others 1993).

  • Are there limits to the applicability of purely passive radon-resistant construction practices?

Essentially no studies have explicitly examined whether there might be an upper limit to the efficacy of purely radon-resistant construction. The research conducted in the FRRP gave some indication that there could be an upper limit. For example, the postconstruction indoor radon concentration was 460 Bq m-3 in a house where the soil-gas radon concentration was 290,000 Bq m-3 (Najafi and others 1995). In another study, the author concluded that radon-resistant techniques could be used in Florida for soil-gas concentrations up to 310,000 Bq m-3, as long as indoor air-exchange rates were kept above about 0.3 h-1 (Hintenlang and others 1994). The modeling done in support of the development of a radon potential map identified areas where the soil radon-potential was high enough that the reduction factors used for radon resistance were not sufficient to ensure that indoor radon concentrations would remain below 150 Bq m-3. These regions were mapped as needing the use of an ASD system, in addition to the radon-resistance features (which often enhance the performance of active systems by limiting air flow between the interior of the house and the depressurized region below the floor slab) (Nielson and others 1994; Rogers and Nielson 1994).

In a few radon-resistant houses examined in the various studies described earlier, indoor radon concentrations exceeded 150 Bq m-3 (based on short-term testing). Most of these houses also had soil-gas concentrations exceeding 40,000 Bq m-3 (where such measurements were done). The degree to which radon-resistance construction standards and guidelines were followed was highly variable in these studies, so it is difficult to determine whether the resulting indoor (basement) radon concentrations above 150 Bq m-3 were due to inadequate construction techniques, low air-exchange rates (in some cases, postconstruction radon testing was done before occupancy), or inherent limits to the principle of radon-resistant construction.

Issues

All the radon or radon-progeny control systems that have been described have failure modes that can reduce or eliminate the effectiveness of mitigation.

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
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What is more, the very nature of a successful installation of a mitigation system is to make it unobtrusive, so that without specifically thinking about it, the building occupant is not likely to know whether the system continues to perform adequately. That is true of mechanical systems—even though pressure gauges and alarms are sometimes used to signal system failure—and it is especially true of passive and radon resistant systems.

One advantage of mechanical systems is that there are objective tests of whether, for example, a fan continues to operate or, when repaired, resumes correct operation. Many ASD systems are installed with pressure gauges of various kinds that provide a necessary (but not always sufficient) measure of continued system performance. Similarly, one can determine that the fan in an HRV system or a filter system designed to remove radon decay products from indoor air continues to operate.

To the extent that radon-control systems also rely on passive radon-resistance techniques to ensure control of radon entry, failures of these features will be much harder to detect without, for example, directly measuring the indoor radon concentration. Even then, the establishment of baseline radon concentrations in a local region is necessary if one is to be able to estimate the overall effectiveness of radon-resistance construction techniques.

System Reliability and Durability

There have been only limited studies of the continued performance of active (that is, mechanically driven) radon-mitigation systems (Naismith 1997; Brodhead 1995; Dehmel and others 1993; Gadsby and Harrje 1991; Prill and others 1990). For the most part, such studies have identified two major sources of system failure: the fan ceases to operate, or is turned off by the building occupant and not restarted. One limitation to these studies has been that relatively few systems have been operating for periods approximating the mean-time-to-failure data for various fans (typically about 10 y). Buildings themselves are expected to last for many decades, if not a century, so multiple failures of a mechanical system should be expected during the lifetime of a building.

Alterations to the building, such as adding ground-contact rooms, can also alter the performance of a radon-mitigation system. In some cases, no provision for additional pipe penetrations is made so the pressure field established by the existing mitigation system does not extend into the region of the addition.

Periodic Radon Testing

One way of ensuring that systems continue to perform adequately is to conduct periodic radon or radon decay-product measurements. Typically, short-duration radon measurements are done following installation of a radon mitigation system in an existing home as a check that the system, as installed, reduces indoor radon

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
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concentrations below the EPA guideline of 150 Bq m-3. Followup testing appears to be very rare. If the building is in a region of the country where radon testing is done as part of real-estate transactions, retesting might occur at that time.

In the case of new construction, particularly where radon-resistance building-construction codes are in place, there is usually no requirement for post-construction testing. Because indoor radon concentrations can be heavily influenced by the operation of a building, such as the use of a heating system (which creates the stack effect), and by occupant behavior, it is essential that radon be measured when the building is occupied. An occupant of a new home built in compliance with a radon-resistance construction code is not likely to have a strong incentive to conduct followup testing. In the case of new construction, the period between completion of construction and occupancy can be several months, and this reduces the likelihood that the construction contractor will have radon testing done.

If control of radon concentrations in indoor air is to be used as an alternative means of reducing radon (decay-product) exposures of the customers of a water-supply system, periodic testing of indoor radon concentrations will be necessary to ensure continued performance of the radon-control methods used. Because these alternatives—reduction of radon concentrations in the drinking water and reduction of radon in indoor air—can be compared only on the basis of health risks (not just indoor radon concentrations), long-term airborne-radon measurements are essential, in that they are the only basis for assessing the health risks associated with airborne radon.

Improved Estimation of the Effect of Radon-Resistant Construction

No rigorous test of the effect of radon-resistance construction practices has been done outside of work done for the FRRP. To be sure, there is evidence that radon-resistance systems can reduce the rate of radon entry into buildings in some cases, but it is not possible to determine the quantitative extent of the reduction on the basis of available data. The data are sparse, both in terms of the numbers of houses examined (for example, the variability in building types and construction practices has not been examined in detail) and in the actual followup testing procedures implemented. It is important to be able to make comparisons between similar houses built in the same area with and without radon-resistance techniques. There are a number of reasons for doing so. Aside from establishing baseline conditions with which to compare the effects of building homes radon-resistant, it will also ensure that the effects can be solidly established. Even though the current EPA building guidelines are applicable mainly in EPA zone 1 areas (the region thought to have the highest radon potential), the vast majority of homes in this region (about 86%) have annual average living-area concentrations below the 150-Bq m -3 guideline. That means that establishing whether radon

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
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resistance is effective will require a careful sampling design to ensure that the comparison can be established with statistical validity.

It is well known that radon concentrations in a house vary from day to day, season to season, and to some extent year to year. Some of the variations are driven by the weather and some by the behavior of the occupants indoors, such as window-opening and door-opening, and furnace and exhaust-fan operation. Thus, such comparisons should be conducted over periods long enough to average out the effects of behavior. Minimum measurement times would be two seasons, but care would have to be taken not to compare data on different houses that were taken during different seasons.

Changes due to the settling and aging of a building substructure over time can create—or extend openings through which soil gas can enter a building. Most radon-resistance approaches use a high-permeability zone just below the slab, created by either a gravel layer or a drainage mat system. The purpose of this zone is to ensure adequate lateral extension of the pressure field created by the passive stack. However, recent theoretical and experimental research has shown that such high permeability zones can substantially enhance radon entry, compared with construction in which the subslab region is not altered (Robinson and Sextro 1995; Gadgil and others 1994; Revzan and Fisk 1992). Not only might the radon-entry rate increase, because this high-permeability zone acts as a uniform-pressure plenum, but the effect of crack size (area) and location is reduced. Even cracks with small total area can transmit as much radon as openings with areas 10—15 times larger (Robinson and Sextro 1995). Thus, it is extremely important to evaluate radon resistance as buildings age to ensure that the radon resistance features are not compromised.

Research and data are needed that would permit reliable estimates of the benefits that might accrue through encouraging the use of radon-resistance construction practices in new houses as an alternative means of reducing radon-related risks. Specific research objectives are discussed in chapter 10.

Mitigation of Radon in Water

One approach to meeting the requirements of the Safe Drinking Water Act Amendments of 1996, with respect to lowering the risk associated with radon, is to treat the water directly. Several studies have evaluated water-treatment technologies for their ability to lower the radon concentration in water. Drago (1998) reported the removal efficiency, flow range, and construction cost of 34 mitigation systems now being used in small and large communities to remove radon from drinking water (table 8.1). The purpose of this section is to present an overview of existing and emerging technologies for removing radon from drinking water.

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
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Table 8.1 Efficiencies, Flow, and Construction Costs for Mitigation Systems Being Used in the United States to Remove Radon from Drinking Water

Treatment Method

Removal Efficiency, %

Flow Range, m3 d-1

Unit Construction Cost, $m-3 d-1

No. of Systems Evaluated

I. Aeration Methods

1. Packed tower (PTA)

79 to >99%

49 to 102,740

18 to 481

11

2. Diffused bubble

a. Single-stage

93

431

312

1

b. Multi-stage

71 to >99

65 to 6,540

11 to 433

8

3. Spray Aeration

~88a (estimated)

1,025

5.3

1

4. Slat tray

86 to 94

1,989 to 2,453

5.3 to 124

6

5. Cascade aeration

~88a (estimated)

5,450

7.9

1

6. Surface aeration

83 to 92a

54,504

42

1

II. Granular Activated Carbon

20 to >99

11 to 981

77 to 365

5

a Estimated.

Source: Drago (1998), Pontius (1998).

Aeration

In July 1991, when EPA proposed regulations for radon in drinking water, it specified aeration as the best available technology to meet the proposed maximum contaminant level (MCL) of 11,000 Bq m-3 . The agency's choice was based on the large removal efficiencies attainable (over 99.9%), the compatibility of aeration with other water-treatment processes, and the availability of aeration technologies in public water supplies. The documentation used to support the decision was published in 1987 (EPA 1987b) and updated in 1988 (EPA 1988a). In the 1991 proposed rule, EPA did not specify a particular type of aeration, but did cite packed-tower aeration (PTA) and diffused-bubble and spray-tower technologies. Mention was also made of less technology-intense aeration methods suitable for small water systems. Evaluations of aeration methods removing radon from drinking water are presented in Lowry and Brandow (1985), Cummins (1987) and Kinner and others (1989).

Aeration methods all exploit the principle that radon is a highly volatile gas and will readily move from water into air. The rate of removal from drinking water is governed by the ratio of the volume of air supplied per unit volume of water treated (A:W), the contact time, the available area for mass transfer, the temperature of the water and air, and the physical chemistry of radon (EPA 1987b). The dimensionless Henry's constant for radon at 20 °C and 1 atm pressure is 4.08 which is higher than values for CO2 or trichloroethylene that are usually removed from water by aeration methods (Drago 1998).

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
×

The differences between the available aeration technologies are primarily a function of the complexity of their design and operation, the flowrates treated, and the radon removal efficiency achieved. The most efficient systems are capable of achieving >99% radon removal by increasing the surface area available for mass transfer of radon from water to air. However, these systems usually require more maintenance than simpler technologies. The more complex technologies are most practical for larger communities that must treat large volumes of water and have a large staff and tax base to support the more extensive capital and operation and maintenance requirements. Most aeration technologies require that the water system operate at atmospheric pressure to allow the release of radon to the air. This means that the systems must be repressurized to supply water to the community. Descriptions of existing and emerging aeration techniques for removing radon from water are given in appendix C.

Issues/Secondary Effects of Aeration

Intermedia Pollution

In its proposed rule, EPA (1991b) recognized that emissions from aeration systems potentially could result in a degradation of air quality and pose some incremental health risk to the general population because of the release of radon to the air. On the basis of EPA's analysis (EPA 1989; EPA 1988b), the increased risk is much smaller than the risk posed by radon in the water. In its initial evaluation with the AIRDOSE model, EPA (1988b) used radon concentrations in water of 68,000 Bq m-3 (range, 37,000–598,000 Bq m-3) based on data from 20 water systems in the United States. Assuming a 100% transfer of radon from the water to air, EPA estimated that radon would be emitted into the ambient air at 0.10 Bq y-1. EPA used an air dispersion model (including radon and its progeny) and assumed ingestion and inhalation exposures in a 50-km radius, and it calculated a maximal lifetime individual risk of 4 × 10-5 (0.016 cancer case y-1). Extrapolated to drinking-water plants throughout the United States, that translated to 0.4 and 0.9 cancer cases per year due to off-gas emissions from all drinking-water supplies meeting MCLs of radon of 7,400 or 37,000 Bq m-3 in water, respectively. EPA used a similar approach to assess the risks associated with dispersion of coal and oil combustion products.

In an evaluation with the MINEDOSE model, EPA (1989) used worst-case scenarios from four treatment facilities whose raw water radon concentrations were 49,000 to 4,074,000 Bq m-3. In only one facility was there a significant potential increase in cancer risk associated with radon emissions when a single point source was assumed. However, EPA found that this large water utility actually used a number of wells at various locations, instead of one source, and that reduced the risk because of dispersion over a greater area. The highest raw water radon concentrations did not always result in the greatest

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
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effect on air quality because high concentrations often occurred in small systems with low flow rates, which yielded lower overall emissions. The evaluation concluded that the resulting health risk posed by radon release into the atmosphere via aeration-system off-gas was much smaller (by a factor of about 100 to 10,000) than the risks that would result if radon were not removed from the water.

The EPA Science Advisory Board (SAB) reviewed EPA's report (EPA 1989; EPA 1988b) and found that the uncertainty analysis needed to be upgraded to lend more scientific credibility to the air-emissions risk assessment. However, the SAB also stated that revisions in the modeling would not change EPA's conclusion that the risk posed by release of radon from a water-treatment facility would be no more than the risk posed by using drinking water that contains radon at 11,000 Bq m-3. The SAB also noted that EPA's assumptions were conservative.

EPA also had its Office of Radiation and Indoor Air (ORIA) review its 1988 (EPA 1988b) and 1989 (EPA 1989) air-emissions studies for consistency and to provide a simple quantitative uncertainty analysis (EPA 1994b). The ORIA review indicated that the early studies had overstated the risk; it estimated an incidence of cancer of 0.004 cases per year, less than the 0.016 case per year initially estimated.

EPA (1991b) acknowledged in the proposed rule for radon that ''some states allow no emissions from PTA systems, regardless of the downwind risks.'' Indeed, on the federal level, EPA has, under the Clean Air Act, established National Emission Standards for Hazardous Air Pollutants (NESHAPs), including radon. Under NESHAP, an average radon-emission rate of 0.74 Bq m-2 s-1 is allowed from radium-containing facilities, and an individual member of the public can have a maximal exposure of 1.00 × 10-4 Gy y-1. It is not clear whether NESHAPs will ever affect aeration systems, inasmuch as they are applicable only to industries, not to drinking-water treatment facilities. Drago (1998) has noted that if the radon NESHAP were applied to a 93 m3 d-1 water-treatment plant (serving about 250 people) with an influent radon concentration of 18,500 Bq m-3 in its water, radon would be emitted in the off-gas at 222 Bq m-2 s-1. This analysis suggests that if the NESHAP were extended to include drinking-water plants, many aeration systems would have to treat the off-gas to remove radon. The Nuclear Regulatory Commission also regulates radon releases from the facilities that it licenses, but it is doubtful that its limits would ever apply to water-treatment facilities. Some states have adopted NESHAPs for radionuclides or Nuclear Regulatory Commission limits by agreement, but none directly applies them to water-treatment plants.

It is possible that some states may conduct or require risk screening of new water treatment aeration facilities for radon emissions. Although some states allow specific incremental lifetime risks associated with hazardous air pollutants (for example, California one in 1 million), it is not clear that they will be applied to radon. Drago (1998) noted that only Nevada, California, New Jersey, and

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
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Pennsylvania have included radon in risk assessments of drinkingwater aeration facilities and that in these cases, radon was part of an evaluation in which volatile organic compounds (VOCs) were the contaminants being regulated, not radon itself.

If radon had to be removed from the off-gas leaving an aeration system, it would present a problem because no efficient and cost-effective technology is available. Granular activated carbon (GAC) is often proposed for off-gas treatment, but as EPA (1991b) acknowledged in its proposed rule, it would probably not be effective in removing radon from air. GAC can remove vapor-phase radon, but it is relatively poorly adsorbed and, according to the equation derived by Strong and Levins (1978), the empty-bed contact time (EBCT) required would be many hours. Martins and Meyers (1993) estimated that less than 2% of radon would be removed by typical vapor-phase GAC units currently used at water-treatment plants to remove VOCs (where EBCT are in seconds). Only with complex methods of concentrating the radon, perhaps by adsorption and desorption (Thomas 1973) and the use of several GAC beds in series could sufficient removal be achieved. Even then, performance would be difficult to control because vapor-phase adsorption is a function of temperature, humidity, and interferences by other gaseous compounds (Bocanegra and Hopke 1987). Indeed, very few studies of techniques to remove radon from off-gas (vapor-phase removal) have been conducted, even in the nuclear industry. Drago evaluated the work of Bendixsen and Buckham (1973) on removing noble gases from gas streams at nuclear facilities (table 8.2.) and found that the methods available are impractical or cost-prohibitive for use in the water industry. As a result, EPA (1991b) suggested in its proposed rule that at sites where air emissions regulations prevent release of radon from aeration systems, GAC might need to be used, instead of aeration, to treat the water. Some other technologies, not known in 1991, when the proposed rule was published, might also be alternatives to aeration, but these are only in the testing phase (for example, vacuum deaeration). The issue of

Table 8.2

Methods to Remove Noble Gases from Gas Streams at Nuclear Facilities

Adsorption

in fluorocarbonsa

in CO2a

on activated carbona

Cryogenic distillationa

Membrane separationa

Cryogenic oxidation with fluorinating compoundb

Cryogenic membrane separationb

aBendixsen and Buckham, (1973).

bDrago (1998).

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
×

radon removal from aeration-system off-gas remains unresolved and would probably be of greatest concern in urban areas.

Airborne release at facilities that treat groundwater can expose operators to high concentrations of radon (Fisher and others 1996). Ironically, the problem has been identified at plants that treat groundwater for contaminants other than radon (such as iron). In a survey of 31 water-treatment plants in Iowa, Fisher and others (1996) found that processes such as filtration, backwashing, and regeneration cause radon release directly into the plant. Their results suggest that the air in all facilities that treat groundwater should be monitored for radon and that ventilation should be investigated as a means of reducing worker exposure.

Microbial and Disinfection-Byproducts Risks

Treated water that leaves aeration systems might contain increased bacterial counts (Kinner and others 1990; 1989). On the basis of the pending Groundwater Disinfection Rule (GWDR) (EPA 1992b), disinfection would be required in cases where the heterotrophic plate count exceeded 500 colony-forming units per milliliter. In addition, aeration systems can have periodic problems with high coliform counts in the treated water as a result of the transfer of bacteria from air to water during treatment. That might also necessitate disinfection to comply with the coliform rule in distribution systems (Drago 1998).

EPA did not mention the potential need for disinfection of the effluent from aeration systems in its 1991 proposed rule, nor did it consider disinfection in its cost estimates for radon treatment. The SAB (1993) criticized the agency for neglecting to do so, and EPA has since added these costs (EPA 1994b). The technology to disinfect groundwater is well developed, and disinfection systems already exist in some communities or are being added to meet the requirements of the pending GWDR. The commonly used disinfection methods include chlorination (such as with sodium hypochlorite or gaseous chlorine) and ultraviolet irradiation.

Although disinfection reduces the health risk resulting from microbial contamination of drinking water, it has its own associated risks, especially if chlorination is used. Groundwater might contain organic carbon at 0.5 to 2 mg L-1 (Cornwell and others 1999; Miller and others 1990; Kinner and others 1990; 1989). Addition of chlorine to water that contains such natural organic matter could result in the formation of disinfection byproducts (DBPs) (that is, trihalomethanes, THMs—such as chloroform, dibromochloromethane, dichlorobromomethane, and bromoform—and other compounds at lower concentrations) in the range of 10–50 µg L-1. THMs are regulated in water under the Disinfection By-Products Rule because they are known to cause cancer in rats and mice. As a result, EPA has established potency values for these compounds. Disinfecting the effluent of aeration systems that remove radon from water increases the risk of exposure to disinfection byproducts. The SAB (1993) criticized EPA for not

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
×

discussing the cancer risks associated with exposure to disinfection byproducts; to date, the agency has not done so.

According to Wallace (1997), finished water produced from surface water tends to have higher THM concentrations than finished water from groundwater supplies. One of the concerns of this committee with regard to methods of reducing radon exposure is the potential for increased exposure to THMs if radon mitigation results in the use of chlorine to disinfect the water to satisfy the pending GWDR. To examine this issue, the committee made a screening level estimate of the relative change in cancer risk associated with surface water and groundwater.

Chloroform concentrations are monitored extensively at water-treatment plants, but only sporadically in residential tapwater. Wallace (1997) has reviewed a number of surveys of water-supply concentrations of THMs. Among them, the Community Water Supply Survey provides representative and comprehensive results (Brass and others 1981). The (average and median) values of chloroform, dibromochloromethane, dichlorobromomethane, and bromoform in supplies derived from surface water were 90 (60), 12 (6.8), 5 (1.5), and 2.1 (<1) µg L-1; respectively, and the average values of chloroform, dibromochloromethane, dichlorobromomethane, and bromoform in supplies derived from groundwater were 8.9, 5.8, 6.6, and 11 µg L-1; respectively, average concentrations all below the detection limit.

The committee used those reported concentrations with unit dose factors for inhalation, ingestion, and dermal uptake and with EPA cancer potencies of the compounds to make approximate risk estimates. The average lifetime cancer risk associated with a surface-water system is around 1 × 10-4 and the corresponding risk associated with a groundwater system is around 5 × 10-5, smaller than the 1 × 10-4 risk of lung cancer posed by inhalation of radon released from water (see chapter 5). The calculations are summarized in appendix D.

Corrosion

Aeration during radon treatment increases the pH of water (Kinner and others 1990; 1989). The increase has been attributed to the removal of CO2 from the water. In a study of aeration units used for VOC treatment, the American Water Works Association (AWWA 1991) reported that the effect of CO2 removal, with the greater stability of CaCO 3 at the higher pH, negated the effect of the increased oxygen concentration in water. There was no increase in the corrosivity of the water. At one very small water-supply system in Colorado, aeration of the water to remove radon actually eliminated the need for addition of lime to prevent corrosion (Tamburini and Habenicht 1992). At a small system in New Hampshire, aeration resulted in a decrease in corrosivity and a reduction in the lead and copper measured in the drinking water (personal communication, D. Chase, Department of Health and Human Services, Bureau of Radiological Health, August

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
×

15, 1997). Although those studies suggest that corrosivity decreases with aeration, it might still be necessary in some systems to add corrosion inhibitors (such as lime and sodium silicate) to reduce the potential for increased release of lead and copper from the plumbing and distribution system.

Precipitate Formations

One problem observed in some aeration systems is formation of precipitates (scale) which can cause operational problems (such as fouling of equipment) and aesthetic concerns (such as release of precipitates to consumers). Depending on the chemical characteristics of the raw water, the most common precipitates are oxide, hydroxide and carbonate species of iron, manganese and calcium. Typical A:W ratios for radon removal systems might be 15:1, and this could result in precipitate formation at iron concentrations as low as 0.3 mg L-1 (Kinner and others 1993).

Common methods of eliminating precipitate-formation problems involve periodic addition of weak acid solutions to clean the equipment or addition of sequestering agents that bind the cations (Dyksen and others 1995; AWWA 1991). Another approach is to install cation-exchange filters before the aeration system. These filters are very effective at trapping iron, manganese, or calcium, but they also concentrate other naturally occurring cations, some of which can be radionuclides (such as Ra2+). The brine used to regenerate the ion-exchange filters can also become contaminated with long-lived radionuclides. In some small-scale applications, aeration equipment is followed by sand or cartridge filters that trap the precipitates (Drago 1998; Malley and others 1993). That is a simple method of removal especially useful in plants that do not have full-time professional operators. After sufficient precipitate has collected in the filter, substantially decreasing water flow, it must be backwashed. Both the brine from the ion-exchange unit and the backwashed material from the sand or cartridge filter are usually discharged to the nearest sewage-treatment system, which in many rural areas is a subsurface leach field. In its cost estimates for radon treatment with aeration, EPA did not adequately consider the cost of precipitate treatment, nor did it address adequately the problems of precipitate formation.

In some cases, it has been shown that iron precipitates have enriched concentrations of long-lived radionuclides, such as radium, lead, or uranium (Cornwell and others 1999; Kinner and others 1990). Depending on the concentrations of these radionuclides, the brine or backwash residuals might require special disposal, as discussed in EPA's (1994c) Suggested Guidelines for Disposal of Drinking Water Treatment Wastes Containing Radioactivity. Special treatment of the residuals can increase the cost of operation and the risk to workers who must oversee the disposal process.

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
×

Granular Activated Carbon

In the 1991 proposed role for radionuclides in drinking water, EPA stated that GAC was not a best available technology for radon removal, although it has been shown to remove radon from drinking water (Kinner and others 1989; Lowry and Lowry 1987; Lowry and Brandow 1985). The agency cited problems with radiation buildup, waste disposal, and contact time. Since then, the SAB (July 1993, EPA-SAB-RAC-93-014 and July 1993, EPA-SAB-DWC-93-015) (EPA-SAB 1993b) has suggested that GAC might be an option for small systems with modest raw-water radon concentrations and that there could be problems with the thoroughness of EPA's analysis of the risk and disposal issues related to the use of GAC. In addition, new data that have become available since 1991 suggest that GAC might require shorter empty-bed contact times than originally thought (Cornwell and others 1999).

GAC was first shown to be an effective technique for removing radon from drinking water in the early 1980s (Lowry and Brandow 1981). As a result of its simplicity, it was installed in many private homes in New England where radon levels were high (1,111,000 to 11,111,000 Bq m-3) (Lowry and others 1991). By the late 1980s, studies of GAC units were producing data that suggested that gamma emissions from 214Bi and 214Pb, the short-lived progeny of radon, were substantial (2 × 10-6-5 × 10-4 Gy h-1) (Kinner and others 1989; Lowry and others 1988; Kinner and others 1987). In addition, accumulation of long-lived species (such as uranium, radium, and especially 210Pb) on the GAC was creating disposal problems (Kinner and others 1990; Kinner and others 1989; Lowry and others 1988). The radon-removal efficiency of some GAC units also decreased with time (Kinner and others 1993; Lowry and others 1991; Kinner and others 1990; 1989). It was the data from the studies in the late 1980s that led EPA to question the use of GAC as a best available technology in its 1991 proposed role. Furthermore, economic evaluations suggested that the cost of GAC treatment is high and in most situations not competitive with aeration, because of the large amount of carbon needed, especially for large radon loadings (high flow or high influent radon concentration) (EPA 1987b). A description of the GAC process is found in appendix C.

Retention of Radionuclides on GAC

By their very nature, GAC systems are designed to sorb and retain contaminants. Thus, while a GAC unit is operating, the bed is accumulating radon. In addition, because of radon's short half-life relative to the GAC unit's run-time (months to years), radon comes to secular equilibrium with its progeny. The solid progeny remain sorbed to the GAC (Cornwell and others 1999; Kinner and others 1990; Kinner and others 1989; Lowry and Lowry 1987) which is not surprising inasmuch as GAC is known to have a high affinity for metals (such as lead)

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
×

relative to the small mass of them created from the decay of radon (Reed and Arunachalam 1994; Rubin and Mercer 1981) (for example, GAC can sorb lead at 6.2 × 104-1.9 × 106 µg kg-1 at a pH of 6.5).

Some groundwater also contains radium and uranium in addition to radon. Uranium can sorb directly to the GAC, but its fate is a function of the pH of the water. At a pH greater than 7, the poorly sorbed, negatively charged carbonate species of uranium, UO2 (CO3)2-2, is predominant. At a pH lower than 6.8, the neutral species, UO2CO3, is predominant and can sorb to GAC (Sorg 1988). Radium is poorly sorbed to GAC (Kinner and others 1990; Clifford 1990; Kinner and others 1989; Sorg and Logsdon 1978) because it forms a hydrophilic species, RaSO4, in water. The pattern and rate of accumulation of uranium, radium, and 210Pb in a GAC unit can be quite different if iron is present. Cornwell and others (1999) found high concentrations of these radionuclides associated with iron-rich backwash residuals from GAC units. That is because radium readily associates with ferric hydroxide and negatively charged metal oxides and hydroxides (Clifford 1990). Uranium also reacts with iron (Clifford 1990; Sorg 1988).

Operational Issues.

During operation of a GAC unit, an equilibrium is established between the radioactivity of radon and its short-lived progeny sorbed to the carbon. The primary problem resulting from retention of radionuclides is worker exposure to gamma emissions from 214Bi and 214Pb. The maximum occupational accumulated dose equivalent per year recommended for radiation workers in the United States is 50 mSv (EPA 1987a). However, EPA has stated that "there is no need to allow" workers in water-treatment facilities that remove naturally occurring radionuclides from water to receive such high annual radiation doses. It further suggests that these workers' annual accumulated dose equivalent should be "well within the levels recommended for the general public" of 1,000 µSv. Hence, EPA has recommended a maximum annual administrative control level of 1,000 µSv until more experience with such situations is gained.

Lowry and others (1988) measured the gamma-exposure fields surrounding 10 point-of-entry units treating water with radon at 96 to 28,074,000 Bq m-3 and achieving removal efficiencies of 83% to over 99%. The gamma exposure rates measured at about 1 m were considerable, in all but one case, because the radon concentration removal was very large (Cnet = 611,000 to 27,926,000 Bq m-3). Except for the 27,926,000-Bq m-3 case, the measurements are in agreement with the range of the calculations from the extended source model.

The gamma exposure to workers can be decreased by using water or lead shields around the GAC units. Lowry and others (1991; 1988) studied the effect of water shielding and lead jackets on the point-of-entry units' gamma exposure fields. For example, at site 9 (table 8.3) (28,074,000 Bq m-3) the maximum gamma-exposure rate at the unit's surface was 73 mR h-1. With a 76.2-cm water shield, this was reduced to 8.0 mR h-1. A 61.0-cm water shield reduced a maximum surface gamma-exposure rate of 4.0 mR h-1 to 0.4 mR h-1 at site 5; and at

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
×

Table 8.3

Maximum Gamma-Exposure Rates and Equivalent Dose Rates from Some point-of-Entry GAC Units at a Distance of 1 ma

Site

Average Radon Cob (Bq m-3)

Average Radon Ctc (Bq m-3)

Exposure Rate (mR hr-1)

Equivalent Dosed (µSv hr-1)

1

97,000

14,000

NAe

NAe

2

613,000

2,200

0.40

4.0

5

1,989,000

165,000

0.186

1.86

5 (with 61-cm water shield)

1,989,000

165,000

0.040

0.40

9

28,104,000

175,000

1.73

17.3

a Lowry and others (1988).

b Input water concentration.

c Treated water concentration.

d Equivalent doses were calculated by the committee and were not included in Lowry and other (1988). See appendix E for method of calculation.

e NA= gamma radiation so low that it was impossible to measure with acceptable accuracy.

about 1 m, there was also a significant reduction. Lead shielding (0.64 cm thick) also reduced gamma exposure.

It is possible to estimate the equivalent dose in millisievert per hour from a GAC unit that removes radon from water if an extended-source model is used (see appendix E). This type of model calculates radiation doses in the vicinity of an extended source of radioactive material including self-shielding. This approach is used in the nuclear industry and in radiation protection to predict the radiation levels associated with a variety of radioactive materials (such as ion exchange units that treat nuclear-reactor cooling water and steam condensate).

GAC units operating in the United States are treating 11-981 m3 of water per day (table 8.1). GAC is also used in point-of-entry applications (water flowrate of 1 m3 d-1). For the purposes of calculating the equivalent dose with the extended source model, the committee used the suggestion made by Rydell and others (1989) that GAC should be used only to treat water that contains radon at less than 185,000 Bq m-3. The unit would be required to meet the MCL, which is assumed to be 25,000 Bq m-3 for this example. (The committee makes no recommendation or endorsement of a specific value for the radon MCL and uses 25,000 Bq m-3 in order to provide a framework for the example.) The results (table 8.4) indicate that as the radon loading (becquerel applied per day) increases with increasing water flow rate over the range of 1 to 981 m3 d-1, the time until the 1,000 µSv maximum equivalent dose is reached decreases from about 7,000 h to about 150 h (1 m from the GAC tank surface). In many cases, it is unlikely that water-treatment plant personnel would need to spend hundreds of hours per year near the GAC units. In fact, the actual number of hours of exposure per worker would be different for each water supply. Certainly, the number of hours of

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
×

Table 8.4

Estimated Equivalent Gamma Dose for Workers at Water-Treatment Plants or in Point-of-Entry Applications Using GACa

Water Flow (m3 d-1)

Estimated Gamma Dose at 1 m (µSv hr-1)

Hours until 1,000 µSv Dose

1 (point of entry)

0.14

7,143

11 (water plant)

0.75

1,333

981 (water plant, pressure driven)

7.0

143

981 (water plant, gravity driven)

6.4

156

a See appendix E for calculations.

exposure would need to be monitored to ensure worker safety. If it were necessary to work on a GAC unit for a substantial number of hours, the gamma emissions could be reduced by first taking the unit off line for about 20 to 30 d to allow the radon responsible for the short-lived progeny to decay.

The calculations also show that as radon loading increases, the dimensions of the tank increase, providing increased absorption of gamma radiation within the tank. Modeling gamma emissions from the tank as a point source will be satisfactory only for very small units (for example, point-of-entry applications). The equivalent gamma dose from a GAC system that removes radon from a public water supply should be modeled with an extended-source model that can be modified to the dimensions of the treatment units. It is clear that treating water that contains more radon (over 185,000 Bq m-3), where high removal efficiencies are required, or at high flow rates (high radon loading) will probably lead to unacceptable equivalent gamma doses to water-treatment plant personnel.

Rydell and Keene (1993) have developed a computer program (CARBDOSE 3.0) that calculates the probable gamma-exposure dose with distance from a typical point-of-entry unit. The program "approximates a 25.4 cm diameter, 12.7 cm high cylindrical volume of GAC as a cylindrically-corrected 24 cm × 24 cm × 13 cm array of 1 cm3 sources using the 72 gamma energies reported for 214Bi and 214Pb and allowing for self absorption and build-up." Rydell and others (1989) reported that the CARBDOSE models' estimated gamma dose rates and the measured values for 10 point-of-entry GAC units were in "reasonably good agreement." They suggest that CARBDOSE can be used as a design tool to estimate the potential gamma radiation exposure during operation of the GAC unit. The committee notes that CARBDOSE should only be applied to GAC units that have very small dimensions (that is, ones that treat very small flows) and are similar to those used in developing the model. Equivalent gamma doses for larger GAC units should be predicted with an extended-source model that can address more-complex geometries.

Disposal Issues.

A few weeks after a GAC unit ceases operation, the major radionuclide remaining sorbed to the carbon is 210Pb because of its relatively long

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
×

half-life (22.3 y). In cases where iron is associated with the GAC or the raw water has a pH less than 6.8, uranium and radium can also be found. These species can pose problems for the long-term disposal of the GAC.

No federal agency currently has legislative authority concerning the disposal of drinking-water treatment-plant residuals that contain naturally occurring radionuclides (Cornwell and others 1999). If the GAC is transported to a site for disposal, the Department of Transportation could regulate its shipment. EPA has published two guidelines that suggest how such wastes might be handled (EPA 1994c; 1990). However, the states are responsible for regulation of naturally occurring radioactive materials (NORM). There have been three detailed reviews of federal and state guidelines and regulations regarding NORM and how they might apply to disposal of GAC used to remove radon from water (Cornwell and others 1999; Drago 1998; McTigue and Cornwell 1994).

The EPA (1994c) guidelines for disposal of water-treatment residuals are centered on the levels of uranium and radium present (for example, in spent GAC or backwash residuals) (table 8.5). Unlike its 1990 draft guidelines, EPA's 1994 version did not cite specific action levels for 210Pb. Instead, because of a lack of conclusive technical data, EPA recommended that the impact of 210Pb contamination be considered case by case. Most states also address NORM wastes on a case by case basis (Drago 1998); the exceptions are Illinois, Wisconsin, and New Hampshire, which have established disposal criteria (Cornwell and others 1999).

The Conference of Radiation Control Program Directors published a draft set of suggested state regulations for technologically enhanced NORM (TENORM), naturally occurring radionuclides whose concentrations have been enhanced by technology (for example by such practices as water treatment). Lead-210 associated with GAC is not specifically addressed in this document, and materials with 226Ra or 228Ra at less than 0.19 Bq g-1 are exempt. The draft recommends flexibility

Table 8.5

EPA Suggested Guidelines for Disposal of Naturally Occurring Radionuclides Associated with Drinking-Water Treatment Residuals

Radionuclide

Bq g-1 (dry weight)

Suggested Disposal Site

Radium

<0.11

Landfill

 

0.11–1.85

Covered landfill

 

1.85–74

Possible RCRA facility (case by case review)

 

>74

Low-level radioactive-waste facility

Uranium

<1.11

Landfill

 

1.11–2.78

Covered landfill

 

2.78–27.8

Possible RCRA facility (case by case review)

 

>27.8

Low-level radioactive-waste facility

210pb

Caution and thorough state-agency review of water treatment and waste disposal plans

 

Source: EPA (1994c).

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
×

in regulating TENORM as long as members of the public receive less than 1 × 10-3 Gy y-1 from all licensed sources (including TENORM).

If spent GAC from a water-treatment plant had enough 210Pb, radium, or uranium associated with it to warrant disposal at either a low-level radioactive waste site or a naturally occurring and accelerator produced radioactive materials (NARM) site, this could have a substantial impact on operation and maintenance costs for the water utility. Actual disposal costs have been estimated as $335 m-3 yr-1 (Kinner and others 1989), approximately $48,000 m-3 (McTigue and Cornwell 1994) and about $11,100 m-3 y-1 (Cornwell and others 1999). In addition, broker and transportation fees would likely be assessed. A typical broker would send trained personnel to the treatment plant to dewater the bed, load and seal the GAC in containers, and decontaminate the site. Cornwell and others (1999) estimated the broker fee at $5,000 (mostly associated with time and travel).

Perhaps the biggest question surrounding GAC disposal is the availability of sites that will accept such radioactive material. Drago (1998) reported that two sites are operating (in Barnwell, SC, and Richland, WA). (Note: Clive, UT, receives only limited low-level and NARM wastes.) However, these facilities are not available to all states. Rather, the Low Level Radioactive Waste Disposal Policy Act (PL-99-240) enacted in 1980 and its amendments (1985) direct states to form compacts with their neighbors and designate a host low-level disposal site. There are nine compacts and one other pending, and five states, Washington, DC, and Puerto Rico are unaffiliated. Low-level disposal sites have been proposed by some of these compacts, but none has been built. The result is that low-level waste generators in all states except North Carolina have access to a disposal facility (Drago 1998), but new facilities are not likely to be readily available in the near future.

Several ways of avoiding the need to dispose of the GAC at a low-level waste facility would not require changing legislation or regulations. Perhaps the easiest would be to dispose of the GAC before radionuclide accumulation necessitates special disposal. McTigue and Cornwell (1994) developed a model that allows operators to predict when a bed is reaching such a level with respect to 210Pb. The CARBDOSE model (Rydell and Keene 1993) makes a similar prediction for POE GAC units. These models are simple to use, and periodic measurements of the actual 210Pb accumulation on the GAC can be made to confirm their estimates. It should be noted that the models do not address the effect of GAC-associated-iron on the 210Pb accumulation (Cornwell and others 1999). If substantial amounts of iron were present in the raw water, such a prediction would be more difficult.

Another alternative to disposal of the spent carbon is thermal regeneration of the GAC that Lowry and others (1990) showed was possible. Both 210Pb and its progeny are volatilized at 850 °C. It is not clear whether release of the 210Pb or 210Po to the atmosphere would be acceptable. If those radionuclides were collected in an air scrubber, they would potentially still present a radioactive-waste disposal problem with respect to the fly ash. Acid regeneration of the spent GAC

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
×

is also possible (Lowry and others 1990). In this case, 210Pb, like stable lead (Reed and Arunachalam 1994), would desorb from the GAC and enter the acid-regenerant solution. However, the spent acid could become a radioactive waste that requires special disposal.

Several authors (Cornwell and others 1999; McTigue and Cornwell 1994) explored the possibility of the GAC's being returned to the vendor (an approach used for GAC used to treat VOCs or substances that impart taste and odor to water). However, the willingness of the manufacturers to do this with radioactively contaminated GAC is not clear, especially for small quantities of GAC (less than 9,100 kg).

The best option overall with respect to disposal appears to be use of GAC in sites where the potential for 210Pb accumulation is minimized (that is, where the radon and iron concentrations in the raw water are low or the water flow rate is low. This would ensure fairly long operating times before the 210Pb reached a critical level likely to necessitate special disposal. The low radon loading would also result in lower risk of worker exposure to gamma radiation.

Long EBCT and High Cost

Bench-, pilot-and full-scale studies of GAC removal of radon have produced estimates of the KSS (adsorption-decay constant, see appendix C) for different carbons (Cornwell and others 1999; Kinner and others 1993; Lowry and others 1991; Lowry and Lowry 1987). Lowry and Lowry (1987) found that the best carbon for radon sorption was a coconut-based GAC (KSS = 3.02 h-1). This carbon has a larger percentage of micropores (0.002 µm) than other types of GAC. It is hypothesized that micropores are most effective for sorbing small molecules and atoms, such as radon gas (Drago 1998).

The cost estimates for GAC treatment have used a KSS of less than 3.02 h-1 (for example, EPA 1987b, KSS = 2.09 h-1). Recent studies by Cornwell and others (1999). specifically designed to calculate KSS values for different carbons, found that for one groundwater with low iron and TOC concentrations, the KSS ranged from 3.5 to 5.2 h-1. These higher values suggest that GAC could be a much more cost-effective option at some sites than originally thought. For example, with a raw-water radon concentration of 111,000 Bq m-3, a flow of 39 m3 d-1, and a KSS of 4.5 h-1, the EBCT and amount of GAC needed to achieve an MCL of 11,000 Bq m-3 (90% removal) would be 31 min and 0.83 m3, respectively, compared with 66 min and 1.8 m3, respectively, if the KSS were 2.09 h-1. Assuming a cost of $883 m-3 of GAC (Cornwell and others 1999), this 54% reduction translates to an $848 savings. However, Cornwell and others (1999) also suggest that a pilot study must be conducted at each site where GAC is being considered, because the KSS will likely be different for each groundwater. For example, for a low-iron, low-TOC groundwater in New Hampshire, a lignite-coal based GAC had an

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
×

average KSS of 4.52 h-1. At a site in New Jersey that also had low iron and low TOC, the average KSS for the same carbon was 2.54 h-1.

Microbial Risk

GAC units often support microbial populations because they provide a good surface for attachment and concentration of organic carbon and nutrients (Camper and others 1987; Graese and others 1987; Camper and others 1986; 1985; Wilcox and others 1983). As a result, GAC units that treat radon have had high concentrations of heterotrophic bacteria in their effluent (Cornwell and others 1999; Kinner and others 1990; 1989). Because heterotrophic plate counts could periodically exceed 500 colony forming units per milliliter, which would violate the proposed GWDR, the treated water would need to be disinfected before distribution (EPA 1992c). Either chlorination or ultraviolet disinfection could be used. Unlike the situation with aeration systems, it is less likely that disinfection byproducts would be formed if chlorination were used after a GAC unit. Disinfection byproducts result from the reaction between chlorine-based disinfectants and naturally occurring organic matter in the water. GAC can sorb the low levels of naturally occurring organic matter in the groundwater (Cornwell and others 1999) until it becomes saturated (Kinner and others 1990). Before saturation, the risk posed by disinfection byproducts would be minimal. Thereafter, it could be similar to that of an aeration system that uses chlorine-based disinfection.

Precipitate Formation

In groundwaters that have high levels of iron, precipitates might accumulate on the top of the GAC bed. This reduces the hydraulic head and contaminant-removal efficiency of the GAC and makes it a poor choice for radon treatment for these types of raw water. Although some iron precipitation has been observed in field evaluations of GAC units that treat radon with low iron (Cornwell and others 1999; Kinner and others 1990; 1989), the problem is usually less acute than in aeration treatment; the water is not usually oxygenated and exposed to atmospheric conditions, so, much less oxidation of the iron occurs. If iron precipitates do form, they present the same problems outlined for aeration systems. Pretreatment to remove the iron before the water enters the GAC is unlikely to reduce the disposal issue unless sequestering agents are used to prevent precipitation. Pretreatment, such as with ion-exchange, would just accumulate the long-lived radionuclides on the resin, also presenting a disposal problem.

In addition, Lowry and others (1990), Lowry and Brandow (1985) and Dixon and Lee (1988) have noted that backwashing releases sorbed radon to treated water, although this has not been observed in all cases (Cornwell and others 1999; Kinner and others 1990; 1989). Desorption of the nongaseous radon progeny has not been observed during backwashing (Lowry and others 1990). In this

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
×

case, disposing of the initial water produced after backwashing and not sending it to the consumer, could eliminate the risk of exposure from release of desorbed radon into the water supply.

Water Storage

Because radon has a relatively short half-life, it is possible to obtain some reduction in concentration by storing the water. It can be stored in separate storage tanks or in those normally used to provide water to a community during periods when demand exceeds the yield of wells. Over 24 h, radon reduction due to storage averages 20–40%, and 5–6 d of storage yields 80–90% losses (Kinner and others 1989). In two waterworks in Sweden, losses of 17–34% were documented during storage (Mjönes 1997). The small reductions mean that this method of treatment is effective only where the percentage removals required to meet the MCL are relatively low and the daily demand for water is small. For example, storage might be an adequate treatment for a school that uses a well to supply water (that is, is not served by a community supply). Repumping is usually required with storage systems used for radon reduction because they are typically operated at atmospheric pressure. Repumping can be avoided if the storage tank is elevated.

Perhaps the biggest problem with this method is providing a reliable and consistent quality of water. As demand fluctuates, the retention time in the storage tank can change, potentially resulting in smaller radon reductions. To avoid that problem, the capacity of storage needs to be increased to ensure acceptable overall radon removal. The tanks usually are vented to the atmosphere, which would increase the risk of air emission as with aeration methods. However, this risk would probably be low because use of storage as a treatment method would be limited to very small water supplies that have relatively low radon concentrations. If radon loss is due solely to decay and not to losses to the atmosphere, there is also the risk of exposure from ingestion of radon progeny such as 210Pb. Because the storage tanks are typically vented to the atmosphere, they might require disinfection under the pending GWDR. As a result, there could be an increased risk of exposure to disinfection byproducts if chlorination is used.

Simple Aeration During Storage

Losses observed after water passes through a storage tank are often higher than those measured when the water resides in the tank undisturbed (Mjönes 1997; Kinner and others 1989). The increased loss has been attributed to aeration that occurs when the water splashes into the tanks. Indeed, the mode of entry is very important. Bottom entry below the water line yields removals similar to loss due to storage alone. Free-fall or entry via spray nozzle or splash box can increase removals to the range of 50–70% (Mose 1993; Kinner and others 1987). Adding a crude coarse bubble-aeration system to the tank can boost removal to the 90%

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
×

range (Kinner and others 1987). The issues associated with these systems are the same as those outlined for storage alone (acceptable percentage removals in spite of variable retention times) and for aeration systems (disinfection byproduct formation upon disinfection, off-gas emissions, and precipitate formation).

Blending

Some public supplies (for example, for some large communities) use a combination of surface water and groundwater. In this case, radon-free or very-low-radon water might be available to mix with the groundwater, which would result in reduction in concentration due to dilution. Blending has been documented as effective in some water supplies (Kennedy/Jenks 1991b; Dixon and Lee 1988), but reductions are usually low (20%) (Drago 1998). The use of blending depends on whether the water is monitored for compliance as it enters the distribution system or when it reaches the first tap (consumer). In the former case, both types of water would need to be available at the same location. In addition, blending can be effective as a best available technology only if mixing is complete.

Reverse Osmosis

In the 1991 proposed rule, EPA specified reverse osmosis (RO) as a best available technology for uranium, radium, and beta-and photon-emitters, but not for radon. RO systems use semipermeable membranes and pressure to separate dissolved species from water. In Sweden, Boox (1995) used an RO filter to treat water in two homes. The systems were installed to improve the taste of the water, but they concurrently reduced the radon content by 90%. RO systems have not been tested for radon removal in the United States, and their use in Sweden exclusively for treatment of radon is doubtful because of their low capacity and relatively high cost. In addition, RO membranes do not work well if turbidity-causing material or precipitates (for example, iron, manganese, silica, and calcium) foul them in the raw water. A brine that contains the contaminants removed from the water is created and must be disposed of. Because the brine is concentrated, the levels of radionuclides might be high, dictating special disposal (as outlined for GAC and aeration systems).

Loss in the Water Distribution System

There have been several studies of this method of radon removal from drinking water. The majority of the decrease in radon results from decay during transit or storage in the closed piping network. Most of the studies have documented losses in the range of 10–20% (Rand and others 1991; Kinner and others 1989; Dixon and Lee 1988). A study in Sweden of four waterworks (Mjönes 1997)

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
×

found higher losses (30–70%), but some of this could have been due to volatilization during pumping or mixing and agitation.

If EPA requires that water meet the MCL when it enters the distribution system, this method of treatment would not be acceptable. Furthermore, it is doubtful that it would consistently produce water with the same radon content, because retention times and therefore losses differ on the basis of the distance of travel and the demand. This method probably would be used only where the radon levels entering the distribution network are relatively low.

Field and others (1998; 1995) have shown that radon levels in a distribution system can actually increase when radon is released during decay of radium associated with iron-based pipe scale. A generation rate would have to be quantified if loss in the distribution system were considered as a treatment alternative in water supplies that have this type of scaling problem. It also has implications for where water samples are collected (at the origin of the distribution system or at the point-of-use) (Field and others 1995).

Vacuum Deaeration and Hollow-Fiber Membrane Systems

These are very new systems that have undergone only laboratory-or pilot-scale testing (Drago 1998; 1997). They have been developed to address the issue of off-gas emissions associated with aeration systems. In both technologies, the radon removed is trapped in a sidestream of water rather than being released to the air. Therefore, these systems have the potential to fill a niche where radon concentrations in the raw water are high (precluding use of GAC) and air emissions of radon are prohibited (precluding use of aeration systems alone). The sidestream water is passed through a GAC bed that sorbs the radon. The GAC is effective in these cases because the radon is dissolved in water, not in a vapor phase. Descriptions of vacuum deaeration and hollow-fiber membrane systems are found in appendix C.

The issues of gamma emissions and disposal of the spent GAC used in the vacuum deaeration and hollow-fiber membrane systems are similar to those for GAC when it is used directly to remove radon from water. Though disinfection might be required to prevent biofilm development on the GAC, microbial and disinfection-byproduct risks are not applicable, because water from the GAC unit used in vacuum deaeration and hollow-fiber membrane systems will not be released to the consumer. The issues of precipitate accumulation and backwashing are also minimized because the sidestream water can be fairly clean. The low transfer efficiency of radon from the sidestream water to the GAC dictates a long EBCT. This, along with the complexity of the systems, would increase the costs of these systems.

Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
×

Conclusions

  • Several aeration methods exist or are being developed to remove radon from drinking water. Aeration was designated as the best available technology by EPA in its 1991 proposed radon rule; however, some issues and secondary effects are associated with these technologies. They discharge radon into the air (intermedia pollution), and the extent to which the off-gas emissions will be regulated is not clear. Removing radon from the off-gas is much more difficult and expensive than removing it from water. Although aeration is, in general, a straightforward technology to use, aeration of groundwater that contains dissolved species, such as iron, can lead to the formation of precipitates (scale) that can cause operational problems, aesthetic concerns, and disposal problems (for example, iron precipitates enriched with long-lived radionuclides).
  • GAC was not listed as a best available technology by EPA in the 1991 proposed radon rule, but it might be an option for small systems that require only minor treatment to meet the MCL. Several issues and secondary effects are associated with GAC. The primary one is retention of radionuclides, which can lead to substantial gamma emissions from the unit and pose a potential problem of radiation exposure for plant operators. Equivalent gamma doses from GAC units that remove radon from public water supplies should be predicted with an extended-source model that can address more-complex geometries. Accumulation of radionuclides can also lead to potential disposal problems for the spent GAC. Both those issues are most problematic when radon loadings are high (that is when treated flows are high or large amounts of radon are retained on the GAC).
  • The treated water leaving aeration and GAC water-treatment systems might require disinfection on the basis of the pending GWDR. If chlorination is used as the disinfection method, trihalomethanes, which are known to cause cancer in rats and mice, could be created. The committee estimated that the average lifetime risk associated with disinfection byproducts formed during chlorination of groundwater is on the order of 5 × 10-5.
  • Storage, blending, loss in the water-distribution system, and other newly emerging technologies (such as vacuum deaeration and hollow-fiber membrane systems) are not likely to be major alternatives for removal of radon from drinking water. Storage, blending, and loss in the water distribution system would be limited to situations where only low removal would be required.
Suggested Citation:"8 Mitigation." National Research Council. 1999. Risk Assessment of Radon in Drinking Water. Washington, DC: The National Academies Press. doi: 10.17226/6287.
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The Safe Drinking Water Act directs the U.S. Environmental Protection Agency (EPA) to regulate the quality of drinking water, including its concentration of radon, an acknowledged carcinogen.

This book presents a valuable synthesis of information about the total inhalation and ingestion risks posed by radon in public drinking water, including comprehensive reviews of data on the transfer of radon from water to indoor air and on outdoor levels of radon in the United States. It also presents a new analysis of a biokinetic model developed to determine the risks posed by ingestion of radon and reviews inhalation risks and the carcinogenesis process. The volume includes scenarios for quantifying the reduction in health risk that might be achieved by a program to reduce public exposure to radon.

Risk Assessment of Radon in Drinking Water, reflecting research and analysis mandated by 1996 amendments to the Safe Drinking Water Act, provides comment on a variety of methods to reduce radon entry into homes and to reduce the concentrations of radon in indoor air and in water. The models, analysis, and reviews of literature contained in this book are intended to provide information that EPA will need to set a new maximum contaminant level, as it is required to do in 2000.

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