Bias in the US Environmental Protection Agency’s Baseline Health Risk Assessment Supporting the Decision to Require Dredging of PCB-Bearing Sediments from the Hudson River
Robert A. Michaels, Uriel M. Oko
CONTINUED FROM PREVIOUS POST
Mobilization of sediment-borne PCBs in dredging
PCB mobilization must be considered in assessing the potential public health significance of PCB dredging. Its consideration by EPA, however, was inadequate. PCB mobilization exacerbated by dredging depends upon three types of cause:
--1. sediment disruption, as by extreme weather events or barge sinkings,
--2. the method of dredging,
--3. and accounting in full rather than in part for PCBs that might be mobilized.
Sediment disruption by extreme weather events. Research undertaken by Joel Baker and colleagues at the Chesapeake Biological Laboratory in Maryland simulating Hudson River PCB dredging (Baker et al. 2001). revealed that EPA modeling lacked the spatial resolution high enough to predict PCB mobilization reliably. They concluded that errors, which could have gone in either direction, probably had in fact underestimated sediment and PCB mobilization from extreme weather events. The authors used this finding to argue in favor of dredging, fearing that harmful PCBs would be mobilized over years if dredging did not remove them. However, removal by dredging presumably also could exert a nearer-term effect episodically.
The method of dredging. EPA’s for the Hudson River PCB site was prepared in 1999 and 2000, when GE planned to dredge hydraulically, via the ‘suction’ method. Indeed, a television commercial campaign by GE impugned the ‘clamshell’ or ‘bucket’ method of dredging as being too dirty. Since preparation of the HRA, however, GE’s proposal has reverted to use of the clamshell method.
Accounting in full for PCBs that might be mobilized. The mass of PCB that will be mobilized may be expressed as a fraction of the inventory of PCB in Hudson River sediments. If the inventory is underestimated, mobilization will be underestimated commensurately. This source of underestimation is addressed with respect to other parameters, below.
In short, EPA’s estimate of PCB mobilization from sediments to the water column and from the water column to the air, together contributing to potential PCB inhalation risks, should not have emerged from EPA's model of the Hudson River; and it cannot materialize in the Hudson River if dredging of PCB-bearing sediments at hotspots actually is undertaken.
PCB congeners to be included in the analysis
All PCB congeners should be included in the inventory of PCBs in Hudson River sediments (Fig. 1). Mono- and dichlorinated PCBs, however, were excluded from the inventory of PCBs in Hudson River water, thereby underestimating waterborne PCBs subject to becoming airborne. Several figures in the revised Hudson River health risk assessment (HRA; US EPA 1999, 2000a, b) depict a precipitous fall-off of“total tri+ PCB congener water column concentrations” within approximately 10 meters of the dredge site. PCB congeners can bind from one to 10 chlorine atoms. If each number of chlorines were represented equally, exclusion of the monochlorinated and dichlorinated PCBs would represent two of 10 (20 percent). The actual fraction (weight-percent) excluded is unclear because commercial PCBs were sold as Aroclors, such as Aroclor 1254, with 54 weight-percent chlorine, such that each Aroclor product sold had a distinctive distribution of mono- to deca- chlorinated PCB congeners (hence the ability to ‘fingerprint’ PCB sources). In addition, PCB degradation in sediments results in gradual dechlorination, which tends to deplete the high-chlorine congeners and enrich the low-chlorine congeners... precisely the congeners that were excluded from the figures, and which apparently were excluded from consideration in quantifying PCB release from river water to air. The fraction of total PCB represented by the monochlorinated and dichlorinated PCBs would appear to be about one third, as suggested by an EPA estimate that is described below.
The plan to dredge Hudson River sediments selected one option from among several remediation options. The option favored by environmentalists, “Alternative no. 5,” would remove 155,000 pounds of PCBs, compared with 1.3 million pounds; which equals 650 tons, or approximately 600,000 kg or 60 tonnes disposed. That is the amount that is reported to have been deposited into the Hudson River by GE from its two upriver capacitor plants before PCBs were banned from U. S. commerce by the Toxic Substances Control Act of 1976. Responding to criticism of the plan to dredge only 100,000 pounds of PCB under a less ambitious option, GE provided ‘new data’ to the EPA that showed that the actual amount of PCBs that would be dredged from the river bottom under Alternative no. 5 would be 150,000 pounds, almost identical to the amount preferred by environmental groups (Cappiello 2001):
“The U. S. EPA says it can dredge 50 percent more PCBs from the Hudson River without increasing the volume of sediment removed” (Cappiello 2001).
By way of explanation, EPA indicated that it simply had refined its PCB estimate of a year earlier. EPA did this by including previously-excluded monochlorinated and dichlorinated PCBs, on the rationale (according to EPA’s TAMS contractor) that “fish principally absorb (higher chlorinated) PCBs.”
The Agency apparently assumed that the monochlorinated and dichlorinated PCBs constituted one third of total PCB (50,000 pounds out of 150,000 pounds of total PCB). Clearly, the Agency’s HRA of 1999 (US EPA 1999) and 2000 (US EPA 2000a, b) for Hudson River dredging therefore excluded approximately one third of total PCBs from the PCB inventory. This was done, notwithstanding that the scope of the Hudson River HRA included the airborne risks, not just fish consumption risks, that might be posed by PCBs that will be resuspended and mobilized by dredging. This exclusion, however, did not stop EPA from taking credit for inclusion in its dredging plan of the extra 50,000 pounds of PCB assumed to be accounted for by the monochlorinated and dichlorinated PCBs to augment the acceptability of its dredging plan in the face of criticism in 2001.
The Agency actions described above highlight three issues relating to potential bias in the scientific analysis:
--1. whether EPA accurately inventoried the amount of PCBs that might pose risks to health,
--2. whether EPA accurately assessed risks potentially posed by PCBs in its PCB inventory (addressed in greater detail later), and
--3. whether the PCB risks that were quantified in the HRA corresponded to the PCB amounts that would be dredged, and that would be subject to mobilization with the potential to pose health risks.
The findings indicate that EPA based its risk estimates on a smaller pool of PCBs. They indicate further that the Agency did this at least in part by excluding monochlorinated and dichlorinated PCB congeners from the HRA. EPA did this, notwithstanding that the excluded congeners would necessarily be included in sediments that would be dredged, and therefore would contribute to airborne PCB concentrations and health risks that might be posed by dredging to people situated near the river. In short, EPA’s estimated PCB residue load contributing to potential PCB inhalation risks should not have emerged from EPA's model of Hudson River and, due to failure to account for mono- and dichlorinated congeners, it cannot materialize in the Hudson River if dredging of PCB-bearing sediments at hotspots is indeed undertaken.
Phases of PCBs to be included in the analysis
All phases should be included, most notably including PCBs that are adsorbed onto particles, molecular PCBs that are dissolved, and particulate PCBs that are colloidal. All PCBs in the HRA, however, were assumed to settle under Stokes Law for spherical silt particles. This assumption constitutes a continuous process of removal of PCBs from the water column, notwithstanding that molecular and dissolved PCB phases would remain because they do not settle. That is, these waterborne PCBs are subject to becoming airborne, but this is not accounted for in EPA's HRA.
The mechanical action of dredging ‘hot spots’ will cause PCBs adsorbed to silt particles to enter the water column. Whereas much if not most of the PCB in the water column will remain adsorbed to silt, a significant, possibly majority fraction will enter the water column in a dissolved (molecular) or a colloidal phase (consisting of microscopic PCB droplets). Exclusion of PCBs in these dissolved and colloidal phases from the revised Hudson River HRA is reported in Appendix E, Section 5.2, titled "TSS Plume Estimates.”In that section, only silt particles were used to estimate settling rates:
"Since data on settling rates were not available, a median value for settling velocity of 1.9 x 10-4 M/sec [16.5 M/d]was used in the transport calculations” (US EPA 2000b).
The above description of settling velocity as a ‘median value’ suggests misleadingly that settling was calculated for a heterogeneous distribution of particles whose median settling velocity is 1.9 x 10-4 M/sec [16.5 M/d]. In fact, only the 'median' value was used. This uniform settling velocity, corresponding to a 20-micron (uM) sphere, excludes dissolved and colloidal PCBs, which are smaller. Dissolved PCBs (bound to water) and colloidal PCBs (subject to Brownian motion and water turbulence) never settle. This unstated simplification overestimates the rate of PCB removal from the modeled water column by assuming that all waterborne PCB is adsorbed to particles that settle at the assumed velocity. Actually, a significant if not predominant fraction of total waterborne (resuspended) PCB will consist of free PCB present in dissolved and colloidal phases.
Inasmuch as silt has specific gravity of about 2.5, the assumed ‘median’ settling velocity corresponds to (spherical) particles of diameter exceeding 20 uM, whereas Stoke’s Law ceases to apply when the settling particles are fines that are less than about 50 uM. EPA's implicitly assumed particle size therefore, also implicitly, assumes that the vastly more numerous PCB molecules in dissolved and colloidal phases will settle at the median rate. Colloidal PCBs are commonly recognized as being 1 uM and smaller and, of course, individual PCB molecules are smaller still. These PCB molecules and colloids also would suspend in the water phase even beyond the dredge site perimeter of perhaps 20 M. Molecular and colloidal PCBs can remain in the water, suspended as globules of pure PCBs that are smaller than 20 uM, without being captured by silt curtains, and without settling at all (Paquin 2001, page 2):
“PCB in colloidal form constitutes the most mobile form of PCB in water, being affected only minimally by settling, physical retention or adsorption. Concentrations of PCB-like compounds in water can be much higher in colloidal form than in suspended solids or in dissolved form, and can be much more difficult to intercept through physico-chemical means” (Paquin 2001, page 2).
Indeed, molecular and colloidal phases of PCB together reasonably may be expected to constitute a significant, possibly the predominant fraction of total PCB in the water column, as illustrated by Table 1. Table 1 shows a site at which dissolved and colloidal PCB together amounted to 54 percent of total waterborne PCB.
An EPA review of experience of dredging PCB shows that dredging hot spots can disperse waterborne PCB beyond a 20-meter envelope ('silt curtain') around a dredge site, with observed concentrations of 0.1 to 0.2 ppm (100 to 200 ug/L, or 100,000 to 200,000 ng/L). This is approximately 3,000 to 6,000 times the PCB concentration assumed under a non-dredging scenario in the HRA prepared in support of another project (specifically, the PSEG NY proposal to site the Bethlehem Energy Center, or BEC, gas-fired power plant on the Hudson River at Bethlehem, New York; Oko and Oko 2001, PSEG NY 2001). In this higher waterborne PCB concentration range, resulting airborne PCB concentrations were reported to have exceeded safe concentrations (24). Indeed, EPA’s HRA Appendix E (US EPA 2000b) states the following:
“While these estimates of total tri+ PCB congener concentrations represent cumulative concentrations, dissolved or particulate tri+ PCB congener concentrations may be of even greater interest. In particular, the dissolved water column concentrations tend to be of greater concern because of their increased bioavailability” (US EPA 2000b, page 59, emphasis added).
In short, EPA’s estimated PCB residue load contributing to potential PCB inhalation risks should not have emerged from EPA's model of the Hudson River and, due to failure to account for dissolved and colloidal phases of PCB in the water column, it cannot materialize in the Hudson River if dredging of PCB-bearing sediments at hotspots is indeed undertaken.
Precipitation of PCB-bearing sediment particles from the water column
Precipitation rates should be quantified realistically, as they in turn quantify the rate of removal from the water column of PCBs that had been resuspended and mobilized by dredging. Instead the residence time of flat, PCB-bearing clay particles in river water was quantified unrealistically, based upon the more rapid precipitation of spherically shaped particles acting in accordance with Stokes Law (Fig. 2). This procedure underestimated waterborne PCBs, and thereby also underestimated the amount of PCB that would become airborne.
Mathematical treatment is simplified when a spherical shape for fine particulates is assumed, which is the case in Stoke’s Law. This assumption, however, predicts faster than natural settling rates because, in nature, spherical particles are rare. Disk, rod shapes, and irregular random shapes are more common, and these shapes settle more slowly than spheres. Mathematical predictions of settling rates that do not account for irregular shapes can predict 100 percent faster settling rate at the >20-uM particle size range, and more than 1,000 percent faster at the <10-um> size range.
Clay is abundant in the Hudson River region, and would constitute a significant if not the preponderant fraction of PCB-contaminated sediment particles that will be resuspended and mobilized during dredging. Flat clay particles settle via a side-to-side oscillation during descent, greatly increasing their path length and residence time in the water column. That is why they settle more slowly than predicted by Stokes Law. Such delay in exiting the water column reasonably would be expected to increase the concentration of PCB-laden particles in the water column markedly, much as delays at highway exits markedly increase traffic on the highway. In short, EPA’s suspended silt cleansing rate should not have emerged from EPA's model of the Hudson River and, due to failure to account for the flatness of clay silt particles, it cannot materialize in the Hudson River if dredging of PCB-bearing sediments at hotspots is indeed undertaken.
Electrostatic charges on PCB-bearing sediment particles in the water column
Clay sediment particles resuspended in water (as by dredging) tend to exhibit negative surface charges. Such particles are maintained in suspension by electrostatic interaction of the negative surface charges with cations (positive ions) in the water column. This electrostatic charge configuration inhibits agglomeration of fine silt particles resuspended by dredging. Electrostatic charges should be accounted for because of their potential importance in inhibiting settling of clay particles and removal of adsorbed PCB from the water column of the Hudson River at dredging sites.
Electrostatic charges should be modeled, but instead they were ignored. By this omission EPA fails to account for prolonged suspension in the water column of charge-separated PCB-bearing clay particles, and it thereby also underestimates waterborne PCBs subject to becoming airborne. Most fine particles, in part because of their high surface-area-to-volume ratio, tend to become electrostatically charged in water (Fig. 3). Clay sediment particles resuspended in water (as by dredging) tend to exhibit negative surface charges. The similar charges cause the particles bearing them to repel one another. The space between charge-separated negatively charged particles then is filled with cations (positive ions) already present in the water column. This configuration of charge separation increases particle residence times in the water column. Some charge-separated particles will not settle at all. Electrostatically separated PCB-bearing particles that do not settle remain in the water column, from which they are more available than settling particles to enter the atmosphere, where they may pose airborne risks.
By excluding this potentially significant factor from the analysis of settling of suspended particles in the Hudson River water column, EPA overestimates the settling velocity of PCB-laden particles to the river bottom, and thereby underestimates the likely concentration of PCBs in the water. In short, EPA’s suspended silt cleansing rate should not have emerged from EPA's model of the Hudson River and, due to failure to account for electrostatic charge separation of suspended silt particles, it cannot materialize in the Hudson River if dredging of PCB-bearing sediments at hotspots is indeed undertaken.
Reflection coefficient of precipitating PCB-bearing sediment particles
The reflection coefficient should be quantified because it constitutes a potentially significant source of return to the water column of PCB-bearing silt particles that are of relatively low mass. If 20 percent of low-mass particles encountering the substrate are swept by currents back into the water column, then EPA's underestimation of the suspended particle population in the water column arising from omitting a reflection coefficient would be 20 percent. We don't know what (if any) single value of the reflection coefficient should be assumed for the Hudson River, or what multiple values might be assumed at each location in the river, under varying flow conditions. Clearly, however, EPA incorporated no reflection coefficient at all (or, equivalently, a reflection coefficient of zero was incorporated) in calculating PCB removal rates from the water column. This procedure thereby underestimated waterborne PCBs subject to becoming airborne.
The rate of free settling in water of silt particles influenced by earth’s gravity can be predicted from particle size and the specific gravity of discrete particles. At the bottom of settling columns where the particles compact, however, other mechanisms take over. One of these processes is reflection (Shavit, Moltchanov and Agnon 2003), which refers to the fact that particles of low mass may bounce off the substrate on which they land. The mass of particles that might be swept back into the water column after settling to the substrate would be expected to be greater in flowing waters, such as the Hudson River, and in laboratory wave chambers (Shavit, Moltchanov and Agnon 2003).
Similarly, colloids may remain in suspension indefinitely as a result of bouncing off water molecules with which they collide (in a well-documented phenomenon termed Brownian motion). The phenomena of reflection and bounce occur in a zone of activity termed the 'hindered zone' of settling. Failure to incorporate a reflection coefficient when calculating settling of PCB-laden particles in the Hudson River water column tends to underestimate particle and PCB concentrations in the water, just as traffic could be underestimated on a highway if the model used fails to count a high fraction of exiting vehicles that immediately reenter the highway. In short, EPA’s suspended silt cleansing rate should not have emerged from EPA's model of the Hudson River and, due to failure to account for reflection of settling silt particles, it cannot materialize in the Hudson River if dredging of PCB-bearing sediments at hotspots is indeed undertaken.
Empirical measurements should be used to validate model assumptions that are made in quantifying PCB entry into the air. Instead, available empirical measurements were diluted with modeled values (see below), thereby underestimating the water-to-air transfer coefficient. Accurate estimation of waterborne PCB entry into the air requires quantification via accounting for PCB codistillation. By ignoring PCB codistillation in quantifying the water-to-air PCB transfer coefficient, EPA underestimated waterborne PCBs subject to becoming airborne. A recent news item (Anonymous 2001) based upon research conducted by the Integrated Atmospheric Deposition Network (IADN 2000) reveals that codistillation has transferred nearly two tons of PCB from Lake Ontario to the atmosphere between 1992 and 1996. According to a news report (Anonymous 2001) describing this startling finding:
"The Great Lakes have begun to 'exhale' significant quantities of chemicals, including ...PCBs..., releasing them into the atmosphere... Researchers say ... the lakes begin naturally cleansing themselves through the volatilization process (i. e., evaporating pollution off the water surface). The latest figures from the Integrated Atmospheric Deposition Network (IADN) show a net release from Lake Ontario alone of almost two tons of PCBs into the air... from 1992 through 1996..." (Anonymous 2001, page 9, emphasis added).
That's a half ton (nearly 500 kg) of PCBs each year codistilling from the surface of a cold lake.Codistillation, however, also is temperature-dependent. Thus it would occur at a greater rate, and to a greater degree, in warm water, such as in Hudson River water that is heated during industrial use as a cooling fluid, and then itself cooled in cooling towers before return to the river. EPA’s failure to account for codistillation might be explained by unfamiliarity with the phenomenon, as well as by an unwillingness to give appropriate credence to empirical data arising from credible reports. In short, EPA’s assumed water-to-air PCB transfer rate should not have emerged from EPA's model of the Hudson River, and it cannot materialize in the Hudson River if dredging of PCB-bearing sediments at hotspots is indeed undertaken.
Empirical measurement of airborne PCBs over PCB-contaminated waters
The degree to which EPA was familiar with PCB codistillation cannot be inferred with certainty. However, such familiarity should have been unnecessary for enabling the Agency to quantify accurately PCB water-to-air transfer coefficients, inasmuch as empirical measurements cited by EPA had been made to quantify them. Indeed, the revised Hudson River HRA (US EPA 2000a, b) Appendix B cites nine empirical measurements of airborne PCB concentrations (Buckley and Tofflemire 1983) contributing toward estimating the transfer coefficient of PCBs from water to air (US EPA 2000b; see EPA’s Table B-1). These and possibly other measurements were used by EPA to produce PCB water-to-air transfer coefficients (summarized in 24; see EPA’s Table B-2; and also see the original data source, Harza 1992) as follows:
“These data can be used to estimate an empirical water to air transfer coefficient, representing the ratio of the PCB concentration in air divided by the PCB concentration in water. Using the detected PCB concentrations in air and water summarized in Table B-2, empirical air-water transfer coefficients range from 0.02 to 0.4 ug/M3 per ug/L, with a median value of 0.09, and an average value of 0.15 (ug/M3 per ug/L)” (US EPA 2000a, page 18).
EPA expressed surprise about the magnitude of these measured values, however, possibly because EPA was unfamiliar with codistillation. In that case the Agency would have expected the transfer coefficients to be lower than those suggested by the measurements. Further investigation could have elucidated the explanation for the higher-than-expected PCB water-to-air transfer coefficients, but further investigation apparently was not undertaken.
Instead, the measured values described above were assigned a low weighting. This EPA accomplished via adulteration of the nine empirically derived transfer coefficients with two lower transfer coefficients that were derived via two modeling approaches (Table 2). The two modeling approaches ignore codistillation, instead producing transfer coefficients consistent with Henry's Law acting on bulk PCB concentrations, that is, assuming even distribution of PCB through water. Model results expressed in units of ng/M2 sec per ng/L could not be compared directly with the empirical values expressed in ug/M3 per ug/L. The units were brought into line, and the comparison made, via use of the average PCB concentration in the river (24 ng/L = 0.024 ug/L; US EPA 2000a, page 18). EPA used this concentration to produce a flux (13 ug/sec; US EPA 2000a, page 19) which, using the median empirical transfer coefficient (0.09), generated a modeled airborne concentration of 0.00012 to 0.00021 ug/M3 (US EPA 2000a, page 20, compared with 0.033 to 0.53 ug/M3 detected empirically (US EPA 2000a, page 20). This corresponds to a factor of 157 to 4,400 difference between the modeled vs. empirical data (0.53/0.00012 = 4,417; 0.033/0.00021 = 157). That is, the modeled water-to-air transport factors downwardly biased the estimated transfer of PCBs from Hudson River water to the atmosphere by a factor ranging from as little as 1/4,400th to 1/157th of the empirically determined values.
EPA’s preference for modeled transfer coefficient values biased the expected concentration of airborne PCBs over the river surface in a direction favorable to EPA’s dredging proposal and, in this sense, this action was self-serving. It was sufficiently self-serving to reduce airborne PCB estimates to below levels of concern to EPA, and below levels of concern to the New York State Department of Environmental Conservation (NYS DEC). Specifically, EPA’s weighting procedure diminished assumed airborne PCB concentrations from above published de minimis levels, requiring quantitative risk assessment, to concentrations below de minimis levels, not requiring quantitative assessment of risks potentially posed by inhalation of mobilized PCBs that might become airborne as a result of dredging (Table 3). Contrary to EPA’s routine procedure of validating its air models against reality via use of dyes or other markers, in this case EPA effectively invalidated empirical data based upon real-world data failing to conform to EPA’s air model. In short, EPA’s estimated water-to-air PCB transfer rate should not have emerged from EPA's model of the Hudson River and, due to failure to adequately consider empirical measurements, it cannot materialize in the Hudson River if dredging of PCB-bearing sediments at hotspots is indeed undertaken.
Warm water sources of Hudson River PCB entry into the atmosphere
Potential warm-water sources of Hudson River PCB entry into the atmosphere, such as cooling towers, must be accounted for in assessing the potential public health significance of airborne PCBs under any dredging scenario. Instead, PCB concentrations resulting from water-to-air transfer were estimated based upon unheated (relatively cold) river water. According to the revised HRA for the Upper Hudson and Mid-Hudson River (US EPA 2000):
"The concentrations of PCBs in air were calculated from a combination of historical monitoring data and modeled emissions from the river…" (US EPA 2000, page ES-4; emphasis added).
For every 10° C rise in temperature, the rate of a chemical reaction, such as the rate of PCB codistillation, may be expected roughly to double. The rate of PCB transfer from water to air occurring with a 40° C water temperature increase accordingly would be expected to undergo four doublings. Thus, the rate at which PCBs in cooling tower water might be expected to escape to the air from water at a temperature of, say, 45° C (113° F) in a cooling tower would be approximately 16 times greater than that in a source of Hudson River water at a temperature of 5° C (41° F).
If dissolved and/or colloidal PCBs rise to 10 ug/L (parts per billion by weight) during dredging, the weight of PCBs entering the cooling tower under one project proposal (the BEC power plant; Oko and Oko 2001, PSEG NY 2001), based on 4,500 gallon/minute uptake, would be 0.25 kg/d (nearly 10 tons/year). Examination of studies forming the basis for the passage quoted above pertaining to transfer of PCBs from river water to air, however, reveal no studies addressing PCB release from warm water in cooling towers. In short, EPA’s assumed low water-to-air PCB transfer rate should not have emerged from EPA's model of the Hudson River and, due to failure to consider warm water sources of PCB entry to the atmosphere, it cannot materialize in the Hudson River if dredging of PCB-bearing sediments at hotspots is indeed undertaken.
Summary of EPA quantification of parameters used in dredging decision making
As documented above, EPA evaluation of the nine subject parameters addressed in this study systematically have underestimated concentrations of PCBs that could, and presumably would, become airborne under non-dredging and dredging scenarios. Adoption of simplifying assumptions in modeling river flow, precipitation of suspended particles, and PCB dynamics can result in omission and/or unreliable quantification of important parameters contributing to overall PCB-associated risk. That this indeed has occurred is hinted at in Section 5 (Assessment of Water Quality Impacts) of Appendix E of EPA’s revised HRA for the Hudson River (US EPA 2000b):
“A complete evaluation of water quality impacts requires integrating a calibrated hydrodynamic model of the system with a water quality model capable of predicting changes due to advection, turbulent diffusion, and settling of the suspended particles. Such a model is beyond the scope of this evaluation” (US EPA 2000b, Section 5, page 12; emphasis added).
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TO BE CONTINUED
Michaels, RA.; and UM Oko. Bias in the US EPA baseline health risk assessment supporting the decision to require dredging of PCB-bearing sediments from the Hudson River. Environmental Practice (Cambridge University Press), 9(2):96-111, June 2007.