Baskaran Report - full version

Executive Summary:

The primary focus of this review is to assess the potential for release of radionuclides from the mixed waste landfill (MWL) and their effect upon human health and the environment. Towards this, a large volume of key documents (publicly available documents) was reviewed. Particularly, the following reports that were carefully evaluated are the following: i) Report of the Mixed Waste Landfill Phase 1 and Phase 2 RCRA Facility Investigation Sandia National Laboratories Albuquerque, New Mexico - September 1991 with appendix; (ii) Response to this report by the Department of Energy (DOE) iii) Report of the Mixed Waste Landfill Phase 2 Facility Investigation Sandia National Laboratories (SNL) Albuquerque, New Mexico, September 1996; (iv) Data obtained from New Mexico Department of Environment; (v) Complete uranium data for the groundwater samples (obtained from SNL). Based on the evaluation of data reported in these reports, it is found that some of the data are ambiguous and high quality data are required to fully characterize the nature and extent of radionuclide contamination. The major issue on the remediation of this site critically hinges on the spatial extent of radionuclide contamination of this site and hence as a top priority, the extent of contamination needs to be characterized.
From a huge database on the radionuclide concentrations in the soil, water, and air in and around the mixed-waste landfill, the following observations have been made:
i) Based on the evaluation of all the data on the concentrations of 235U and 238U and activity ratios of 235U/238U in the groundwater samples, the data appears to indicate that a significant portion of U in the groundwater samples has a non-natural source (s).
ii) The activity ratios of 238Pu and 239,240Pu in the bore hole soil samples collected below the MWL and analyzed by three different groups indicate that there is some amount of Pu observed in the bore hole soil samples and it is likely derived from non-global fallout sources.
iii) Based on the activity ratios of 238Pu and 239,240Pu in the aerosol samples collected at the MWL, it appears that most of the Pu is derived from non-global fallout sources.
iv) The tritium has migrated far and wide from the MWL and it continues to migrate to adjoining areas.
v) Although modeling efforts predicted that the radionuclides will not migrate more that 30-40 ft from the MWL, presence of non-natural U, Pu and 90Sr at >100 ft below ground surface indicates that the models are not well constrained.
vi) The concentrations of any of these nuclides are below the upper tolerance limit and have not posed any serious threat to the environment so far.
It must be made clear that the concentrations of these radionuclides in the groundwater system at present do not pose any threat to the environment. Due to the quality of the radiochemical data presented in the reports, the uncertainties associated with measurements are unusually high and it is not possible to unequivocally conclude if non-natural uranium or Pu derived from the mixed waste landfill has reached the groundwater system. High quality data on the uranium activity ratios in soil, water and air samples from the MWLF will provide information on the sources of radionuclides. The radon emanation rate measurements as well concentration of 222Rn in the air samples show that there is no significant amount of radon input from the daughter products of 238U buried in the MWLF.
Various multiple remediation scenarios have been evaluated. Those include: i) No-action ii) Containment and iii) Some combination of retrieval, treatment, storage, and disposal. For options ii) and iii), the relative merits of the following techniques have been evaluated: a) Capping to prevent mobilization of contaminants; b) In-situ treatment; c) Dynamic compaction; d) Soil Grouting; In-situ vitrification; e) soil washing/chemical extraction and any other methods that may be relevant to this site. Some information on the costs involved for remediating this MWL have also been discussed.

I. Evaluation of the Risk Assessment Strategy Proposed by the Department of Energy
I.1 Introduction:
Contamination of land and water by chemical and radioactive materials derived from the contaminated nuclear sites is a subject that has caused much public concern. Several methods have been adopted towards lessening the contamination or its impact by removal, treatment, or stabilization (this process is known as 'environmental remediation' or 'environmental restoration').
There are several steps involved in the remediation process for any site (Eisenbud and Gesell, 1997). In this particular case, there are four major phases of the remediation process.
i) Preliminary assessment and site investigation
ii) Maintenance and Surveillance
iii) Remedial investigation and feasibility study
iv) Corrective measures
The first step in the remediation process for any site is to make an inventory of the contaminants and the hazard. Based on the evaluation of the content and hazard, one of the following three options can be chosen for remediating that particular site. The three basic options for contaminated sites are: I) No action; ii) Containment; and iii) some combination of retrieval, treatment, storage, and disposal (DOE, 1989).
Based on a detailed investigation of various sets of reports generated on this particular mixed waste landfill, the following observations have been made:
i) 6,300 Ci of radionuclides (Low-level Waste (LLRW) and Mixed Wastes) have been dumped in the Landfill.
ii) The classified area contains the mixed waste where longer-lived radionuclides such as Pu-238, Pu-239,240, etc are present.
iii) Some of the concrete caps used in the MWL have already worn out and a clear assessment of the capped material needs to be done; in addition, in the event of land subsidence, the potential possibilities for contaminant release from these pits and trenches need to be investigated.
iv) Between September 1990 and October 1995, several rounds of quarterly and semi-annual groundwater sampling have been conducted. Out of these, few analyses of radionuclides were conducted. The uranium concentrations in the groundwater samples range from 0.0028 to 0.0078 mg/L, although concentration in MW-4 well (2.69 mg/L) is about 100 times higher than the proposed EPA drinking water standard of 0.020 mg/L and this value is discarded without any scientific reasoning. The report concludes that the uranium detected in groundwater at the MWL is most likely from natural sources. The 238U/235U activity ratios reported in the report appear to indicate presence of non-natural uranium in the groundwater.
v) Although 90Sr concentrations in the MWL groundwater are below the DOE DCG (derived concentration guide, DCG, 40 pCi/L), concentrations ranging from 2.2 to 5.7 pCi/L were detected in groundwater samples. Again, doubts were raised on the quality of data and the DOE discarded this data set.
vi) In 1967, 271,000 gallons of reactor coolant water was disposed in Trench D. Assuming a residual volumetric soil moisture content of 9%, and a surface area of 4,500 ft2 for this trench, it was estimated that the total depth of penetration would be less than 100 ft. This is a conservative estimate under the assumptions that there is no lateral spreading and evaporation.
vii) Model prediction (BOSS) on the contaminant flow and transport suggests the following: a) Cesium and strontium will only migrate to a depth of less than 33 feet from the point where they were released; b) no detectable activities of tritium in groundwater were predicted (although detectable levels of tritium was reported); and c) no detectable radionuclide activities in groundwater is likely to occur now or in the future (although several radionuclides, such 90Sr, U have been reported).
viii) Vapor-phase transport is a more significant transport mechanism than aqueous-phase transport at the MWL, because tritium and ppb levels of volatile organic compounds (VOCS) have migrated from the disposal pits and trenches. However, this mode of transport will not explain the migration of U and 90Sr to the water table.

I.2 Environmental Setting:
The nearest Meteorological station to the mixed MWL is at the Albuquerque International Airport, which is about 6 km from the MWL. The average annual rainfall is 8.4 inches of which about 50% occurs during intense summer thunderstorms between July and September. Potential evapotranspiration for this area is about 31 inches. Actual evaporation has been estimated at 95% of the annual rainfall (Thomson and Smith, 1985, Johnson et al., 1995). Although the average annual wind speed for the Albuquerque area is 9 miles/hour, very high wind speeds during spring times are common. The radioactive and chemical wastes are estimated to extend to a depth of 25 feet.

I.3 Inventory of Nuclear Wastes (estimated to be as of 1995):
Classification Volume of Waste Amount of Radioactivity
Transuranic Waste (TRU) 50-600 ft3 > 1.7 Ci*

Fission-Product Wastes 12,000 ft3 200 Ci (1995, mainly 90Sr & 137Cs)

Uranium & Thorium wastes 22,000 ft3 >10 Ci

Induced activity wastes 54,000 ft3 3,500 Ci (Co-60, Ni-63, and Ba-133; all other have likely decayed away due to short half-lives)

Tritium containing waste 12,000 ft3 2,400 Ci

Low-level wastes 2,600 ft3 6 Ci (mainly alpha-emitting, low-
level Alpha-emitting wastes)

*: Based on the inventory of radionuclides in the MWLF; ~20 microspheres of 238Pu with a size range of 2 to 20 mm were buried in Pit SP-2; 0.1 g of 238Pu (~1.7 Ci) was buried in Pit 21.

II Assessment of Earlier Data
II.1 Groundwater
Groundwater monitoring has been conducted since 1990. Four wells (three down-gradient and one background) were installed and these wells have been sampled quarterly during the first 18 months and then annually thereafter. In none of the samples, levels above the DOE Derived Concentration Guide are found. However, the claim by the DOE that extensive analyses of radionuclides have demonstrated conclusively that there is no groundwater contamination to date at the MWL is highly questionable. It is very well known that the activity ratios of 238U/235U in natural samples is a constant and can be calculated from the natural abundance of these two nuclides and their decay constants [238U/235U activity ratio = (natural abundance of 238U/natural abundance of 235U) * (decay constant of 238U/decay constant of 235U); 238U/235U AR = (0.992745/0.00720)*(7.04 x 108 /4.46 x 109)=21.76]. From the known values, the expected 238U/235U ratio in any natural sample is 21.76. In all four groundwater samples as well as the adjoining background sample, this ratio is significantly less than 21.67 (Table 1). In all the 31 water samples for which uranium analysis was done, the uranium concentrations and 238U/235U activity ratios are given in Appendix-A. In some cases, the activity ratio is considerably lower than 21.67. This clearly indicates that there is at least another non-natural source U and that is likely to have been derived from the mixed-waste landfill (either from the coolant water that was discharged which reached the groundwater or other mechanisms by which U migrated and reached the groundwater. It is pertinent to point out there is some (although the quality of data is not highly reliable due to very high uncertainties, similar to the soil samples) Pu and 90Sr are also present which could be attributed to the same source as U. In addition, large amounts of depleted uranium as well as 235U (Pit SP-1) have been buried.
In addition to the field data, considerable amount of modeling work has been done on this MWL to predict if there would be any migration of radionuclides within the next 30 years. Models presented in the report by Johnson et al. (1995) as well as Klavetter analysis indicate that detectable activity in the groundwater is not likely now or in the next 30 years while U activity ratios and presence of small amounts of 90Sr and Pu suggest that these nuclides somehow reached the groundwater already. It appears that the parameters used in the models are not well constrained, as there seems to be a discrepancy between the field data and the model prediction.

Table 1: Concentrations* of 238U, 235U, 90Sr, and 239,240Pu in Groundwater samples Collected in October 1995

Sample Code(Well depth in ft.) 238U (pCi/L) 235U (pCi/L) 238U/235U (AR)** 239,230Pu (pCi/L) 90Sr (pCi/L)
MW-1 (478) 2.23±0.25 0.176±0.065 12.7±4.9 - 0.15±0.30
MW-2 (477) 2.26±0.22 0.169±0.054 13.4±4.5 0.028±0.024 -
MW-3 (476) 1.86±0.20 0.146±0.050 12.7±4.6 0.03±0.31 -
MW-4 (Upper, 548) 1.76±0.19 0.116±0.048 15.2±6.5 0.003±0.012 0.16±0.31
BW-1 (477) 2.21±0.22 0.187±0.055 11.8±3.7 0.01±0.014 0.30±0.32

-: Unreliable data, either due to 0 or negative values
* The uncertainties are 2 sigma uncertainties given in the report
** The 2-sigma uncertainties in the concentrations are propagated.

II.2 Soil:
II.2.1 Summary of data on tritium in soil samples:
Soil samples (both surface and at depths) around the MWL to assess the radionuclide transport from the MWL have been routinely collected for various chemical and radiochemical analysis since 1969. These samples were commonly analyzed for gross beta activity, 90Sr, 137Cs, tritium, and uranium and gamma activity. From the samples collected in 1969, Brewer et al (1973) concluded that there was no evidence of radionuclide migration from the MWL. According the Johnson et al. (1995), the raw data for this study and the exact sample locations are not available.
In 1979, one pit was dug to a depth of 35 feet adjacent to classified area. Samples were collected at one-foot intervals. Results indicate that uranium and gross alpha measurements were found to be below detection limit. However, there are two observations that show some migration of radionuclides. Gross beta activity in the samples from a depth of 29 feet was 10 times higher than the mean of the total number of samples. In 25 out of 35 samples, tritium levels were higher than the background values (Johnson et al. 1995). Tritium concentration in the extracted water at a depth of 23 feet was found to be 389,000 pCi/L. The raw data along with the exact location of the sampling pit are also unavailable.
In 1981, 226 surface and subsurface soil samples were collected at 10 locations in the MWL area
and at one location at the perimeter of the landfill (used as a background location). Tritium levels were below detection limit (< 450 pCi/L) at the background location. Tritium concentrations in surface soil outside the disposal area and borings within the disposal area ranged from 1,000 to 139,000 pCi/L. The highest tritium concentration (3.51 x 108 pCi/L) was detected in surface soil located adjacent to Pit 33 in the classified area which was active at that time (Johnson et al., 1995). It is pertinent to point out that this concentration is about 1000 times higher than the maximum value reported in the samples collected in 1979. According to Johnson et al. (1995), the original data along with the exact location of the sampling points are unavailable.
From the measurements of tritium on one hundred samples collected in 1982 to determine the lateral extent of tritium migration in and around the unclassified and classified area (especially Pit 33), maximum tritium concentration of 3.3 x 106 pCi/L was found at a site adjacent to Pit 33. It was estimated that the total inventory of tritium in the soil within the classified area to be 1.18 x 109 pCi, of which 96% was estimated to be in the area adjacent to Pit 33. Additional investigations from the soil borings to a maximum depth of 90 feet adjacent to Pit 33 and outside the classified area indicated that no detectable levels of tritium were found outside the boundaries of the classified area at any depth. The original data along with the exact locations for this study seems to be unavailable.
In 1989, during installation of groundwater monitoring wells down gradient from the landfill, soil samples were collected. Tritium analyses on these samples indicated that the concentrations were below detection limit (Johnson et al., 1995). This suggests that significant amounts of tritium had not migrated to areas outside the periphery of the MWL.
Ten surface samples out of a total of 17 on which tritium analysis was carried out in 1990 had tritium levels between 7,000 and 390,000 pCi/L. Assuming that the detection limit is similar (i.e., 450 pCi/L) to earlier measurements, these values are highly elevated. Several subsurface samples were collected (including few samples from beneath the landfill) and analyzed for gross alpha, gross beta, tritium, selected gamma emitters, and atomic ratios of U (235U/238U) and Pu (239Pu/240Pu). Results of radionuclide analysis indicate that the concentrations were below detection limit and there is no migration of any radionuclide, except tritium. Tritium was detected in 14 of the 18 bores collected and in the remaining 4 samples it was below 2,000 pCi/L, the detection limit (it is not clear why this detection limit is different from 450 pCi/L reported for the samples collected in 1981). The highest concentration, 1.7 x 107 pCi/L, was observed at a depth of 15 ft in bore SB-3. From this set of samples, it is clear that lateral migration of tritium from the classified area to outside the fenced area has taken place (Johnson et al., 1995). In one bore, the detectable tritium concentration was observed at a depth of 110 feet (the MWL pit is typically about 25 feet deep), which implies that the tritium has migrated about 85 feet through the vapor phase.
In 1993, subsurface samples were collected for tritium during slant boring drilling operation beneath the MWL, with entry just north of Trench D. The tritium concentrations at depths >35 ft ranged between 100 and 870 pCi/L extracted water. At 394-400 ft, the tritium level was 760 pCi/L. Surface samples had the highest concentrations, up to 1.0 x 106 pCi/L. This data agree with the earlier set of data and clearly show that subsurface migration of tritium has taken place and the contamination has extended out beyond the MWL line.
More recent data on the soil samples on-site, KAFB perimeter, and Community soil indicate that there is considerable scatter on the tritium data. For example, in on-site location #11, the tritium activity varied between below detection limit to 390 pCi/L (on water extracted from soils). In August 1995, the concentration was 20 pCi/L while in August 1997, it increased to 190 pCi/L. However, similar trends are not seen uniformly in all the samples. Surface soil sample tritium results are given as Appendix-D.
In summary, migration of tritium beyond the burial site has taken place. In the 24 years time period during which tritium was monitored, tritium has migrated to considerable distances. For example, if the burial depth of a pit is 25 feet, it has migrated all the way to 400 ft over this time period. Although the inventory of tritium is about half as that of 1988 when the MWL was closed, tritium continues to migrate to farther distances from the MWL.

II.2.2 Plutonium contamination in the soil:
Pu concentration was measured in a few soil and air samples and the concentrations have been reported in Phase-2 Report (1996). Some of the soil samples were re-analyzed by the SNL and NMED for Pu. NMED got re-analyzed four soil samples, one from each depth at 50', 140', 499' and 546'. Presence of Pu at deeper depths (as deep as 499 feet) has major ramifications and thus, it is important that the Pu data obtained is analyzed rigorously.
In order to compare the quality of Pu data, all the data reported in Phase-2 Report, samples reanalyzed by SNL as well as by NMED are reported in Table 2. In Phase-2 Report, significant Pu concentration has been reported at depths as far as 496 feet below ground surface (0.06 pCi/g). It is not clear why the standard deviation is so high (0.06 ± 0.11 pCi/g). In the re-analysis of 4 soil samples by the NMED (analysis done at the American Environmental Network Inc.), the concentrations are lower than those reported in the Phase-2 Report. In the re-analysis of some of the samples by the SNL, 238Pu concentrations are significantly higher than the MDL values (reported by the Core Laboratories where the analytical work was carried out; CORE LAB, 1998). For example, 238Pu concentration in the soil sample collected at 499 feet below surface was reported to be 1.03±0.72 pCi/g while at 546 ft, the concentration was found to be 0.68±0.35 pCi/g. From the three data sets analyzed on the same set of samples and the propagated uncertainties associated with them, it is not possible to unequivocally conclude if Pu is present at soils 499 and 546 ft from the ground surface. The major impediment is the data quality generated by these three groups.
If the data is real, it might appear to be a mystery how Pu got that far, unless Pu-bearing fluid was discharged and the fluid penetrated all the way to ~500 feet below ground surface. It would be highly desirable to have the activity ratios 238U/235U in the soil samples at various depths, as that would provide clear evidence on the sources of uranium in the soil (more discussion in the groundwater section). It is pertinent to give a perspective on the quality of Pu data here. When we had analyzed Pu in soil/sediment samples, the precision is about 10 to 100 times better than the data presented in this report and thus, we could clearly identify the sources of Pu in the environment (for example, Baskaran et al., 1995, 1996, 2000). In order to substantiate this point, the data reported in Phase-2 report as well as some of our data are given below (in both cases, 2 sigma uncertainties are shown; the concentration of Pu-239, 240 only given since Pu-238 concentration in our soil/sediment samples are very low with Pu-238/Pu-239, 240 activity ratios around 0.03 while in the aerosol samples, the concentrations appear to be high, clearly indicating possibly non-global fallout source(s)):

Arctic soil/sediment data
from Baskaran et al., 1996) Phase-2 Report data
3-R 0.28±0.11 pCi/kg ER92003652-3 50 ft 10±20 pCi/kg
64B (5-8cm) 0.34±0.19 pCi/kg ER92003655-4 70 ft 60±30 pCi/kg
10 0.98±0.24 pCi/kg ER92004043-3 89 ft 10±20 pCi/kg
40 1.19±0.42 pCi/kg ER92004041-3 199 ft 2.5±7.6 pCi/kg
8A 1.67±0.49 pCi/kg ER92004348-3 499 ft 60±110 pCi/kg

The uncertainty associated with the Pu and U measurements lead to several questions:
i) If the Pu concentrations reported for depths 199, 499 and 546 feet below ground surface are real, how did Pu get there? Since Pu is highly particle-reactive (in oxidized and reduced forms; all valence states, although oxidized state has a lower particle affinity for particles compared to the reduced state species), it is unlikely this movement is due to movement of meteoric water.
ii) If the data is real, then, it is likely that the aquifer is also contaminated with Pu. With this type of quality of data, it would be very difficult to conclude whether Pu is present in the aquifer or not.
iii) Of all the five soil samples on which Pu concentration was determined and reported in Phase-2 report, only one sample had significant (non-zero value with 2 sigma) 238Pu concentration (ER92003655-4, 70 feet below surface ground level) of 50±30 pCi/kg (239,240Pu = 60±30 pCi/kg), with 238Pu/239,240Pu activity ratios of 0.83±0.65. This value is distinctly different than the global fallout value of 0.03 for the Northern hemisphere (e.g., Baskaran et al., 1995, 1996). If we take all the data generated on the re-analysis of samples by SNL, 238Pu is present in all the soil samples, all the way to 546 ft below ground (all values are above MDL). If this Pu is one of the sources for the aerosols at this site, then, the 238Pu/239,240Pu activity ratio in the aerosol could be different than the global fallout value of 0.03.

II.3 Radionuclide Measurements on Air Samples, Flux Measurements of Tritium and Radon and Radiation Survey Study in the Mixed Waste Landfill:
Air sampling was conducted and analyzed for gross alpha and gross beta measurements during the years 1989-1990, 1992 and 1993. In most of the aerosol samples, the gross alpha and beta activities were below detection limit. The concentrations of 239,240Pu and 238Pu in the aerosol samples are given in Table 3. The activity ratios of 238Pu/239,240Pu in the global fallout for the northern hemisphere are 0.030 (Baskaran et al., 1995, 1996). Although the 2-sigma uncertainties associated with the measurements are very high (and hence the propagated uncertainties on the activity ratios), most of the Pu in the air samples appears to have been derived from non-global fallout sources. If some of the Pu is derived from the MWLF or adjoining areas with 238Pu/239,240Pu ratios different than the global fallout value, then, other radionuclide measurements could also provide information on the sources of those radionuclides. For example, if there is some non-natural uranium (with 238U/235U activity ratio distinctly different than 21.76) is present in the air, then, the activity ratio of uranium will provide information on the source of uranium. Since activity ratios of U (238U/235U) are not reported, it is not possible to conclude anything (whether the source of U is natural or derived from the MWL) on the sources of uranium.

Table 2: Concentrations* of 238Pu, and 239,240Pu in soil samples collected at the MWLF

Sample Code(Well depth in ft.) Data reported 238Pu (pCi/g) 239,240Pu (pCi/g) 238Pu/239.240Pu (AR)**
MW-4 (50) Phase-2 ReportReanalysis-SNLReanalysis-NMED 0.04±0.040.67±0.420.00±0.03 0.01±0.020.160±0.2700.00±0.02 4.0±8.94.2±7.6-
MW-4 (70) Phase-2 ReportReanalysis-SNLDuplicateReanalysis-NMED 0.05±0.030.900±0.6400.520±0.420NA 0.06±0.030.00±0.010.00±0.01NA 0.83±0.78---
MW-4 (89) Phase-2 ReportReanalysis-SNLReanalysis-NMED 0.03±0.040.830±0.440NA 0.01±0.020.110±0.240NA 15.2±6.57.5±16.8-
MW-4 (140) Phase-2 ReportReanalysis-SNLReanalysis-NMED 0.003±0.0090.700±0.580NA NR0.00±0.010NA ---
MW-4 (200) Phase-2 ReportReanalysis-SNLReanalysis-NMED 0.003±0.0140.610±0.420NA 0.0025±0.00760.230±0.300NA 1.2±6.71.5±2.2-
MW-4 (499) Phase-2 ReportReanalysis-SNLReanalysis-NMED 0.01±0.111.03±0.720.02±0.02 0.06±0.110.00±0.010.00±0.02 0.17±1.9--
MW-4 (546) Phase-2 ReportReanalysis-SNLReanalysis-NMED NR0.680±0.350-0.01±0.03 0.064±0.0720.00±0.010.00±0.03 ---
NR: Not reported
NA: Not analyzed
* The uncertainties are 2 sigma uncertainties given in the reports
** The 2-sigma uncertainties in the concentrations are propagated.
Tritium flux measurements were made across the MWL. Measured fluxes of tritium varied between 41 and 166,000 pCi/m2/hr (Johnson et al., 1995). A strong correlation between the tritium concentrations in the soil samples and tritium flux was observed (Johnson et al., 1995). Radon flux testing was completed on the Mixed Waste and Classified Waste Landfills, Technical Area III, SNL/NM. The radon flux measurement technique employed for this study was capable of detecting radon fluxes in the range of 1-2% of the 20 pCi/m2/s limit listed in draft 10 CFR 834. The mean flux (range: 0.026 to 1.021 pCi/m2/s) value is slightly below the background radon flux value of 0.45 pCi/m2/s for soil listed in NCRP 97, Measurement of Radon and Radon Daughters in Air (Radon Gas Survey, 1998).
Radiation levels at the classified area were significantly higher than those in the unclassified area. At the west central border of the classified area, radiation level of 5 mR/hr was reported (Johnson et al., 1995). This value is considerably higher than the background value for the Albuquerque area. Buried wastes in pits in the classified area resulted in this high level of radiation.

Table 3: Concentrations* of 238Pu and 239,240Pu in aerosol samples collected in October 1995 (Table 4.2-1 PM10 sampling results, Phase-2 Report, 1996)

Site 238Pu (x 10-4pCi/L) 239,240Pu (x 10 -4pCi/L) 238Pu/239.240Pu (AR)**
MWL East Side 4.6±2.7 2.7±1.9 1.7±1.5
MWL West Side 19.0±0.57 2.5±2.0 7.6±6.1
MWL Upwind 1.2±0.92 2.2±1.6 0.55±0.58
* The uncertainties are 2 sigma uncertainties given in the report
** The 2-sigma uncertainties in the concentrations are propagated.

III.1 Mechanisms of Transport of Radionuclides:
There are five potentially significant release and transport mechanisms that can be identified for this MWL. They include: i) Subsurface movement of radionuclides from the MWL to the groundwater system and subsequent transport through the groundwater movement; ii) Advection and volatilization of radionuclides in the gaseous phase; iii) Disturbance of this site by biota (burrowing, grazing, etc.); iv) uptake of radionuclides by plants and organisms and subsequent movement through the food chain; and v) soil erosion leading to the exposure of buried mixed waste and subsequent mobility of radionuclides by wind (by saltation of sedimentary particles, and resuspension of very fine particles), rainfall (mainly via runoff), and biota. The importance of each of these release and transport mechanisms is discussed below.
Subsurface movement of radionuclides from the MWL can take place by one or more of the following major routes: i) Infiltration of and percolation of rainfall through the unsaturated zone to the water table; and ii) migration of radionuclides through the vapor/gaseous phase. Of all the radionuclides that were dumped, there are only two nuclides that can migrate through the gaseous or vapor phase, viz., tritium and radon-222. The mobility of all other radionuclides is primarily through water movement. The distribution coefficient (Kd = concentration of a nuclides in the solid particles (pCi/g) / concentration of the nuclide in the solution phase (pCi/cm3)) of tritium is very low as compared most other radionuclides. From the analysis of samples collected in 1990, it has been documented that tritium has migrated about 85 feet (= 110-25 ft) over a period of 2-20 years. From 1990 till present, it is likely that it has moved further down. It is conceivable that further movement of tritium and significant amount even could reach the water table
It is very likely at the present rainfall amount, there is not going to be any significant movement of all other radionuclides due to percolation of rainwater, except tritium and radon. It is also likely that even if all the containers decompose completely, the migration of radionuclides due to the percolation of precipitation is unlikely. However, the fact that there is non-natural uranium in the groundwater samples and potentially some Pu present in the soil samples all the way to ~500 feet below the ground surface with 238Pu/239,240Pu activity ratios very different than the global fallout values, clearly indicate there is a contribution of some of the non-gaseous radionuclide contribution from the MWL to the groundwater. In addition, if present day climate is only half or a third as wet as the long-term average, and the area goes through periods with three times as much moisture, then, these factors could lead to migration of some of the radionuclides stored in the MWL.
Although the subsurface movement of nuclear contaminants from the present MWL is unlikely, there could be other mechanisms that could lead to the migration of some of these radionuclides (especially those ones that are soluble, such as Sr-90 and to some extent Cs-137). One such mechanism that one can envision is the tectonic subsidence of a block and subsequent transport of water through fast-flow paths in the soil. Recently, in one of the most extensively studied sites for the burial of nuclear wastes (Yucca mountain site), Fabryke-Martin et al. (lead author from Los Alamos, and some of the researchers are from Lawrence Livermore National Lab) showed that fast transport of water via fractures to subsurface. Although the geology and geohydrology of the Yucca mountain site (annual rainfall is only 6 inches) is very different than the MWL in Albuquerque, the potential possibility of faster percolation of water through the fractures cannot be ruled out. In addition, there are some 'three sigma' and 'four sigma' events that we need to be prepared for. The possible scenarios are:
i) If there is 20-50 inches of snowfall in a year at the site and slowly the water melts and all the water seeps through the ground. This could correspond to 5-13 inches of water, which could reach the dumped wastes. If all the waste containers are decomposed, then, radionuclides such as Sr-90 and Cs-137 can migrate relatively easily. If such snowfall continues for a few years, then, there could be some serious problems.
ii) If there is some tectonic subsidence taking place and the block that contains the mixed wastes undergoes subsidence, then, potentially water can seep through and again reach the contaminants. This could also potentially lead to migration of radionuclides (as well as other organic contaminants) and eventually they could reach the groundwater table.
iii) If there is any major earthquake at the MWL, then, there could be serious problems and probability that these radionuclides can get into the groundwater could be significantly higher than at present. The danger in this case is twofold- rainwater flowing into the pits and trenches from above and water being pushed up from an aquifer 300 feet below by an earthquake.
The second factor is the advection and volatilization of radionuclides in the gaseous phase. Only gaseous/vapor phase radionuclides can be transported through this pathway. Although, transport can occur through either liquid or gaseous phase, the pertinent pathway for this MWL is through the vapor phase. There are only two radionuclides that could potentially migrate through this pathway: radon and tritium. Since tritium behaves chemically similar to hydrogen (being an isotope of hydrogen, only physical properties such as density, vapor pressure, boiling/melting point, etc., are different), when tritium interacts with water, it becomes a part of the water molecule (1H3HO) and can migrate as water vapor or liquid soil water. As discussed before, tritium has migrated from the buried to surface soils as well as 85' below the mixed waste burial depth.
Although disturbance by plants and animals may not be the major mechanism, in rare circumstances, this can potentially pose serious problems. For example, Prairie dogs are common in and around this area, and they could burrow into shallow areas. This activity could lead to the accelerated migration of radionuclides such as tritium and radon.
Uptake by plants (through their roots) and organisms could lead to the transport of radionuclides from this site to other places. Since there are no trees or plants (except some grasses) at the MWL, it is likely that plants may not play a major role in the transport of radionuclides. Animal ingestion of contaminants is another transport mechanisms, but no attempt has been made to quantify this aspect.
Erosion of soil could lead to the exposure of buried contaminated waste, if the capping is degraded.

III.2 Can the Reactor Coolant Water have reached the groundwater and potentially contaminated the soil and groundwater?
The 271,000 gallons of reactor coolant water disposed in 1967 in Trench D may have contained several radionuclides (Johnson et al., 1995). The reactor coolant water contained ~ 1 Ci total radioactivity, mainly short-lived activation products. This trench may have already contained waste at the time of discharge of the coolant water and subsequently some of the radionuclides could have migrated through the water movement. If 271,000 gallons of coolant water was discharged, it could potentially reach all the way to the groundwater table. For example, if it was discharged at 25 ft from surface at a surface area of 200 ft2, the total volume of soil in 425 ft column is 85,000 ft3 (635,800 gallons). If the porosity is 25%, then, the total volume required filling the pore space equals 158,960 gallons of fluid. This is a simple calculation and assumes that the fluid was discharged at an area of 200 ft2 and the fluid moves vertically downward and there is no evaporation of the fluid. It does indicate potentially the fluid can reach the groundwater table.

IV. Other Potential Solutions and Considerations Towards Removing the Nuclear waste from the Landfill and Dumping Elsewhere:
There are three basic options towards remediating the contaminated mixed waste landfill sites. They are: i) No action; ii) Containment; and iii) Some combination of retrieval, treatment, storage and disposal. Based on the results and interpretation made from Phase-1 and 2 studies, the Department of Energy has selected the capping alternative towards this landfill. When risk assessment clearly demonstrates that there are no unacceptable risks associated with leaving the contaminants in place, leaving the waste in the burial site is a better option.
Various options have been proposed for remediating the contaminants in the mixed waste landfill. Those include: a) Capping the contaminants; b) in-situ treatment and dynamic compaction; c) soil grouting; d) soil washing/chemical extraction. The relative merits and demerits are discussed below.
a) Capping: Capping prevents the mobilization of contaminants by precipitation and collection of water that could percolate through or run off the site.
b) In-situ Treatment: In situ treatment can be performed within the contaminated material at the site. For example, dynamic compaction can be used to reduce voids and volumes of waste. However, any liquid present in containers in the MWL could potentially rupture during the treatment and the liquid could move further down, even to the saturated aquifer zone. In addition, the waste will basically remain in the same place. In situ vitrification is another experimental process in which electrodes are inserted into waste material and current is passed though the electrodes to heat the material (Jacobs and Spalding, 1988; Alexiades et al., 1994). At relatively high temperature, volatile toxic organic compounds will be driven off (including some amounts of tritium). The residual material is collected and melted in place to form a glasslike substance that retains the nonvolatile contaminants in a stable form. In this process, the volume of the waste is reduced, but the nuclear contaminants will remain in tact (except tritium), and hence a site to bury this vitrified material needs to be identified. This may not be the best solution for this MWL
c) Soil Grouting: For the grouting purpose, a stabilizing material (such as polyacrylamide) is mixed with the waste and the mixture is injected into the wells where the material flows into voids and fractures, hardens, and immobilizes the waste. This basically results in addition of other chemicals and the only benefit is the reduction of the volumes of the waste. In addition, the amount of radioactivity will remain the same.
d) Soil Washing: This is a treatment process that removes contaminants from excavated soils using physical separation, chemical extraction, or a combination of the two (Parikh et al., 1995). Physical separation uses screening and water to separate soils into several size fractions while chemical soil washing is removal of contaminants by mixing contaminated soils and extractants in a continuous leaching process. This method is more suitable for a contaminated site and it is not the best method for the mixed waste landfill.
Based on the merits and demerits of these 4 options, it appears that no-action or removal of the waste and dumping it elsewhere (at least partial removal) would be the options for this landfill. The major factors on digging out and dumping the mixed-level waste in another place are: i) High radiation due to high activity at present, which could be a health hazard for the workers; ii) Costs involved with the digging out and dumping them elsewhere; and iii) Lack of appropriate site for the dumping of this waste. These three issues are discussed separately below.
There is no question that the amount of dosage one could receive from the removal of waste could be excessively high. Due to relatively short-ranges of alpha and beta emitters, the alpha and beta particles can be attenuated easily. The only major problem is the gamma-emitting radionuclides, such as 60Co. For example, 1-Ci source of 60Co handled at a distance of 50-cm (without any shield) could lead to a radiation dose of 9 rem/hr (Appendix-B). However, if the handling distance is increased to 5 meter, then, the radiation dose received will decrease by a factor of 100, to 90 mrem/hr. It is very likely that the DOE has the capability to handle this issue, as operations through Robots have been conducted elsewhere (for example, mixed waste landfill site in Idaho site; it is quite conceivable that the DOE has the capability to handle these and other high-level wastes in sites like Savannah River site and Hanford site). It would be difficult to sort the radioactive material from the volatile chemicals and other toxic contaminants. If there is a mechanism to sort out the volatile chemicals, then, intense heat can be applied to solidify into solid material resembling obsidian that can be stored elsewhere.
If DOE has spent 10 million $ over the last 11-12 years (from 1991 to 1996, 5.2 million $ have been spent on MWL characterization) since the mixed waste dumping was stopped in the landfill, then, by investing some more money we can solve the problem once for all. Assuming that the DOE will continue to spend money towards maintenance and surveillance in the next few decades, it may even be economical to spend the needed money to dig this waste and dump it in another nuclear waste site, at a reduced cost. It may even be worthwhile to consider excavating those pits and trenches where the activities are significantly higher [such as Pit-33; from 1959 to 1983, 1861 Ci of 3H was dumped of which 1451 Ci was discharged in the classified area. Of this amount 822 Ci was dumped in Pit 33 (between May 1979 and January 1983)]. This would require a complete updated inventory of various radionuclides in all the pits and trenches and then decide which ones need immediate removal.
The third major issue for the DOE is where do we put the waste, if we dig them out. There is at least one commercial company located in Utah, which works with DOE closely. The website for that company is: http://www.envirocareutah.com/main.html. Potentially, this company can accept some of the wastes from the MWL site.

V. Recommendations:
Any remediation process must first start with the assessment of the extent of contamination and to make an inventory of the contaminants and the hazard. It is the evaluation of this consultant that the assessment on the extent of contamination is not fully characterized. The following questions that remain unanswered need to be addressed. The questions and what possible steps can be taken are given below. This consultant is fully aware of the additional costs involved for the analysis of samples, but it would be few thousand $ as compared to the 5 million $ spent over the years from 1991 to 1995 for the characterization of this site. It would be very important that the analysis is done in ultra-high quality lab (preferably in academia where the scientists will be fully conscious of each number they generate and can defend the quality of the data)
i) Question to be Answered: Is there any non-natural source of uranium to the groundwater in the Mixed Waste Landfill?
Possible analysis: Collect 2 water samples from each of the four wells + 1 background well and measure concentrations of 235U and 238U by thermal ionization mass spectrometry (TIMS). The required sample size for TIMS is small; however, if the analysis is to be done by alpha spectrometry, then, sample size should be increased to 5-liters and the detector in alpha spectrometer should have very good resolution and very low background in the regions of interest. The uncertainty on the 235U concentration should be less than 5%.

ii) Question to be Answered: Is there any non-natural source of uranium to the soils at various depths in the Mixed Waste Landfill?
Possible analysis: Since soil samples have already been collected from MWL, those samples can be utilized for this analysis. In all the layers where Pu analysis was done, 238U and 235U analysis should also be done. It is preferable to measure the concentrations of uranium by TIMS. However, if the analysis is to be done by alpha spectrometry, then, sample size should be 2-3 grams and the detector in alpha spectrometer should have very good resolution and very low background in the regions of interest. The uncertainty on the 235U concentration should be less than 5%.

iii) Question to be Answered: Is there any plutonium to the groundwater in the Mixed Waste Landfill?
Possible analysis: Collect 2 water samples from each of the four wells + 1 background well and measure concentrations of 238Pu and 239,240Pu by alpha spectrometry. The sample size should be 10-20 liters and the detector in alpha spectrometer should have very good resolution and very low background in the regions of interest. The uncertainty on the 238Pu concentration should be less than 5%.

iv) Question to be Answered: Is there any plutonium in the soil samples at various depths in the Mixed Waste Landfill?
Possible analysis: Since soil samples have already been collected from MWL, those samples can be utilized for this analysis. In all the layers where U analysis was done, 238Pu and 239,240Pu analysis should also be done. If the analysis is to be done by alpha spectrometry, then, sample size should be 2-3 grams and the detector in alpha spectrometer should have very good resolution and very low background in the regions of interest. The uncertainty on the 238Pu concentration should be less than 5%.

v) Question to be Answered: Is there any non-natural source of uranium to the aerosols near in the Mixed Waste Landfill?
Possible analysis: Collect 2 aerosol samples from the MWL site and another from farther away (background sample) and measure concentrations of 235U and 238U by thermal ionization mass spectrometry (TIMS). If the analysis is to be done by alpha spectrometry, then, large volumes of air must be filtered. The detector in the alpha spectrometer should have very good resolution and very low background in the regions of interest. The uncertainty on the 235U concentration should be less than 5%.

vi) Question to be Answered: Is there any Pu source derived from the MWL to the aerosols near in the Mixed Waste Landfill?
Possible analysis: Collect 2 aerosol samples from the MWL site and another from farther away (background sample) and measure concentrations of 238Pu and 239,240Pu alpha spectrometry. The detector in the alpha spectrometer should have very good resolution and very low background in the regions of interest. The uncertainty on the 235U concentration should be less than 5%.

References:
Alexiades, V., G. K. Jacobs, and N. W. Dunbar. 1994. Constraints on mass balance of soil moisture during in situ vitrification. Environ. Geol. 23, 83-88.
Baskaran, M., S. Asbill, P. H. Santschi, T. Davis, J. M. Brooks, M. A. Champ, V. Makeyev and V. Khlebovich. 1995. Distribution of 239,240Pu and 238Pu concentrations in sediments from the Ob and Yenisey Rivers and the Kara Sea. Applied Radiation and Isotopes 46, 1109-1119.
Baskaran, M., S. Asbill, P. H. Santschi, J. M. Brooks, M. A. Champ, D. Adkinson, M. R. Colmer, and V. Makeyev. 1996. Pu, 137Cs, and excess 210Pb in Russian Arctic sediment. Earth and Planetary Science Letters 140, 243-257.
Baskaran, M., S. Asbill, J. Schwantes, P. H. Santschi, M. A. Champ, J. M. Brooks, D. Adkinson, and V. Makeyev. 2000. Concentrations of 137Cs, 239,240Pu and 210Pb in sediment samples from the Pechora Sea and biological samples from the Ob, Yenisey Rivers and Kara Sea. Marine Pollution Bulletin (in press).
Blush, S. M. and T. H. Heitman (1995) Train wreck along the river of money-An evaluation of the Hanford cleanup. U.S. Senate Committee on Energy and Natural Resources, Washington, D.C.
Brewer, L. W. (1973) Environmental Monitoring Report for Sandia Laboratories from 1964 through 1972.
DOE (1989) Environmental restoration and waste management five-year plan. Report DOE/S-0070. U.S. Department of Energy, Washington, D.C.
Eisenbud, M. and T. Gesell (1997) Environmental Radioactivity, 4th Edition, Academic Press.
Friedlander, G., J. W. Kennedy, E. S. Macias, and J. M. Miller. (1981) Nuclear and Radiochemistry, 3rd edition, Wiley Interscience.
Jacobs, G. K., and B. P. Balding. 1988. In situ vitrification demonstration for the stabilization of buried wastes at Oak Ridge National Laboratory. Nucl. Chem. Waste Manage. 8, 249-259.
Johnson, R., D. Blunt, D. Tomasko, H. Hartman, and A. Chan, 1995. A Preliminary Human Health Risk Assessment for the Mixed Waste Landfill, Sandia National Laboratories, Albuquerque, New Mexico, Argonne National Laboratory, Argonne, Illinois.
Parikh, S. R., R. D. Belden, and K. E. Cook, 1995. Pilot-scale soil washing treatability test to investigate the removal of radionuclides from contaminated soils in the 100 Area of the Hanford Site. In: Proceedings of the Symposium "Environmental Restoration 95," August 13-17 in Denver, Colorado, U.S. Department of Energy Office of Environmental Restoration, Washington, D. C.
Radon Gas Survey, 1998. Appendix-E- Sandia National Laboratories, January 19, 1998.
Thomson, B. M. and G. J. Smith. 1985. "Investigation of Groundwater Contamination Potential at Sandia National Laboratories, Albuquerque, NM" in Proceedings of the Fifth DOE Environmental Protection Information Meeting, held in Albuquerque, NM, 6-8 November 1984, p. 531-540, CONF-841187.

Appendix: B
Calculation of Dosage Rate in rads per hour to be expected at a distance of 50 cm from 1-Ci 60Co source:
Rad is a quantitative measure of radiation energy absorption (usually called the dose). A dose of 1 rad deposits 100 ergs g-1 of material.
Each disintegration of 60Co is accompanied by two g quanta with energies 1.17 and 1.33 MeV; for simplicity we take the average of these energies for calculation (=1.25 MeV).
1 Ci = 3.7 x 1010 disintegrations per second. Number of quanta emitted = 3.7 x 1010 emissions per second x 2 quanta/emission = 7.4 x 1010 quanta per second.
At a distance of 50 cm, the g flux is = 7.4 x 1010 quanta/s / [4p x 50 cm x 50cm] = 2.3 x 106 photons cm-2 s-1
This corresponding gamma ray energy is calculated as follows:
3.2 x 106 photons cm-2 s-1 x 1.25 x 106 eV/photon = 2.9 x 1012 eV cm-2 s-1
The fractional energy loss for the gamma rays per g cm-2 of air is given by a factor 0.055 (this is obtained from the relationship m/r = 0.693 g cm-2/12.5 g cm-2 where 0.693 is a constant and 12.5 g cm-2 is the half-thickness for air from established data). The energy lost by the g rays in going through 1 g cm-2 of air is
0.055 x 2.9 x 1012 (eV cm-2 s-1) x 3600 (s/h) = 5.7 x 1014 eV hr-1 or 920 erg hr-1. If we set the energy absorbed per gram of air equal to this energy loss, we get 920 erg hr-1/100 = 9.2 rad hr-1.
Quite frequently, the unit of radiation dosage that is used in radiation protection is the roentgen equivalent man (rem). Dosage in rems = Dosage in rads x relative damage caused by various radiation
Rems = Rads x Quality Factor (QF)
For gamma radiation, QF = 1
Thus, radiation dosage received due to 1 Ci of 60Co at a distance of 50 cm is = 9.2 rem hr-1. This can be compared to the values set for the population at large as well as individuals whose occupation entails exposure to radiation. For those who are occupationally exposed it is recommended that the whole body radiation not exceed 5 rems yr-1 (Friedlander et al., 1981). The recommended exposure in addition to background and medical procedures for the general population is 0.50 rem yr-1.

Glossary:
DOE: Department of Energy
DCG: Derived concentration guide for water (DOE, 1990)
Kd: Distribution coefficient (a measure of the affinity of a nuclide to particle)
MWL: Mixed Waste Landfill
NMED: New Mexico Environment Department
Mixed Waste: It is a waste that contains a hazardous waste component and a radioactive material component. The hazardous waste is either listed under 40 CFR Part 261, Subpart D, and/or exhibits a characteristic described in 40 CFR Part 261, Subpart C. Hazardous waste component is regulated by EPA. The radioactive material must be classified as one of the following: source, special nuclear or by product material subject to the Atomic Energy Act of 1954 (AEA) (42 U.S.C. Section 201). The DOE regulates radioactive material from the mixed waste in DOE facilities.
PCi: pico curie (10-12 Curie; 1 Curie = 3.7 x 1010 disintegrations per second)

Appendix-A - U-238 and U-235 concentrations and activity ratios in Groundwater Samples

Sample Collection U-238 Error-2 s U-235 Error-2 s U-238/U-235 Error
Date (pCi/L) pCi/L (pCi/L) (pCi/L) Activity Ratio 2-s propag.
MWL-BW1 1-Sep-90 2.38 0.0629 37.8 NO DATA
MWL-BW1 1-Sep-90 2.43 0.0782 31.1 NO DATA
MWL-BW1 1-Sep-90 2.36 0.0833 28.3 NO DATA
MWL-BW1 1-Sep-90 2.28 0.0909 25.1 NO DATA
MWL-BW1 10-Nov-93 3 0.5 0.071 0.064 42.3 38.7
MWL-BW1 10-Nov-93 3 0.41 0.11 0.066 27.3 16.8
MWL-BW1 27-Oct-94 3.18 0.81 0.26 0.19 12.2 9.5
MWL-BW1 27-Oct-94 3.05 0.8 0.43 0.26 7.1 4.7
MWL-BW1 23-Oct-95 2.21 0.22 0.187 0.055 11.8 3.7
MWL-MW1 3-May-94 1.8 0.42 0.094 0.062 19.1 13.4
MWL-MW1 4-May-94 1.9 0.42 0.1 0.06 19.0 12.1
MWL-MW1 25-Oct-94 2.45 0.64 0.38 0.22 6.4 4.1
MWL-MW1 20-Oct-95 2.23 0.25 0.176 0.065 12.7 4.9
MWL-MW2 8-Nov-93 2.6 0.44 0.12 0.082 21.7 15.3
MWL-MW2 2-May-94 2.2 0.49 0.23 0.1 9.6 4.7
MWL-MW2 24-Oct-94 2.33 0.7 0.2 0.18 11.7 11.1
MWL-MW2 17-Apr-95 2.41 0.25 0.184 0.062 13.1 4.6
MWL-MW2 16-Oct-95 2.26 0.22 0.169 0.054 13.4 4.5
MWL-MW3 9-Nov-93 2.5 0.39 0.099 0.066 25.3 17.3
MWL-MW3 3-May-94 2 0.46 0.089 0.064 22.5 17.0
MWL-MW3 25-Oct-94 2.2 0.71 0.32 0.25 6.9 5.8
MWL-MW3 17-Apr-95 2.02 0.22 0.163 0.058 12.4 4.6
MWL-MW3 16-Oct-95 1.86 0.2 0.146 0.05 12.7 4.6
MWL-MW4 11-Nov-93 2.4 0.21 0.13 0.042 18.5 6.2
MWL-MW4 14-Mar-94 2.8 0.49 0.21 0.082 13.3 5.7
MWL-MW4 31-May-94 2.1 0.48 0.061 0.048 34.4 28.2
MWL-MW4 28-Oct-94 2.94 0.85 0.22 0.2 13.4 12.7
MWL-MW4 19-Apr-95 1.81 0.22 0.171 0.062 10.6 4.0
MWL-MW4 19-Apr-95 1.81 0.21 0.142 0.056 12.7 5.2
MWL-MW4 20-Oct-95 1.72 0.2 0.139 0.051 12.4 4.8
MWL-MW4 20-Oct-95 1.76 0.19 0.116 .048 15.2 6.5