Journal of Hazardous Materials 174 (2010) 648–655

Contents lists available at ScienceDirect

Journal of Hazardous Materials
journal homepage: www.elsevier.com/locate/jhazmat

The effects of preferential ﬂow and soil texture on risk assessmentsof a NORM waste disposal site
E.M. Pontedeiro a , M.Th. van Genuchten b,∗ , R.M. Cotta b , J. Simunek cabcBrazilian Nuclear Energy Commission, CNEN, Rua General Severiano 90, Rio de Janeiro, RJ 22290-901, Brazil
Department of Mechanical Engineering, COPPE/LTTC, Federal University of Rio de Janeiro, UFRJ, Rio de Janeiro, RJ 21945-970, Brazil
Department of Environmental Sciences, University of California, Riverside, CA 92521, USAa r t i c l e

i n f oArticle history:
Received 2 August 2009
Accepted 18 September 2009
Available online 9 October 2009
Keywords:
Radionuclide transport
NORM wastes
Preferential ﬂow
Risk assessment
Solid waste disposala b s t r a c t
This paper investigates the environmental fate of radionuclide decay chains (specially the 238 U and 232 Th
series) being released from a conventional mining installation processing ore containing natural occurring radioactive materials (NORMs). Contaminated waste at the site is being disposed off in an industrial
landﬁll on top of a base of earth material to ensure integrity of the deposit over relatively long geologic
times (thousands of years). Brazilian regulations, like those of many other countries, require a performance assessment of the disposal facility using a leaching and off-site transport scenario. We used for thispurpose the HYDRUS-1D software package to predict long-term radionuclide transport vertically throughboth the landﬁll and the underlying unsaturated zone, and then laterally in groundwater. We assumed
that a downgradient well intercepting groundwater was the only source of water for a resident farmer,
and that all contaminated water from the well was somehow used in the biosphere. The risk assessment
was carried out for both a best-case scenario assuming equilibrium transport in a ﬁne-textured (clay)
subsurface, and a worst-case scenario involving preferential ﬂow through a more coarse-textured subsurface. Results show that preferential ﬂow and soil texture both can have a major effect on the results,
depending upon the speciﬁc radionuclide involved.
© 2009 Published by Elsevier B.V.

1. Introduction
Much attention over the years has focused on the possible longterm effects of waste disposal on the environment and humanhealth. Numerical models can be important tools for landﬁll
designers and operators, as well as for environmental regulators,
to predict the long-term subsurface transport of contaminantsleached from solid waste disposal sites as a function of climatologic
conditions and landﬁll design. Performance assessment is especially important for landﬁlls containing radioactive mining wastes
resulting from the processing of naturally occurring radioactive
materials (NORM and TENORM wastes), often involving the U and
Th radionuclide decay series [1–3]. TENORM refers to technologically enhanced naturally occurring radioactive material in whichhuman activities (e.g., mining of ores or drilling of oil and gas) haveincreased, in comparison to the altered state, the concentration ofnaturally occurring radionuclides, and/or enhanced the potential
for radiation exposure to humans and biota [4].
A large number of models are now available to predict the transport of aqueous contaminants from solid waste landﬁlls or similar∗ Corresponding author.
E-mail address: rvangenuchten@yahoo.com (M.Th. van Genuchten).
0304-3894/$ – see front matter © 2009 Published by Elsevier B.V.
doi:10.1016/j.jhazmat.2009.09.100

facilities in soils and groundwater. They range from relatively simple mass balance models designed speciﬁcally for landﬁlls, suchas the HELP code [5], to more comprehensive multi-dimensionalmodels for more general variably-saturated ﬂow and contaminanttransport problems, such as the TOUGH [6] and HYDRUS [7] families
of codes. While these models have seen application to a wide rangeof problems, including for performance studies of waste disposalsites and capillary barriers, they have been used only sparingly in
combination with detailed risk assessments of NORM waste sites[1,8].
Safety or dose assessments for long time periods are essential
for the management of long-lived NORM and TENORM contaminants. Since institutional controls may not last much longer thana few decades after site closure, the wastes are unlikely to remain
safe. Natural leaching of radionuclides into the adjacent biosphereis almost certainly to occur within a few centuries or even earlier.
Since NORM wastes in Brazil are typically deposited in industrial
landﬁlls, regulations require a safety assessment of the disposal
facility using a leaching and off-site transport scenario.
In this paper we evaluate radionuclide transport from a NORM
disposal site in Amazonia using the HYDRUS-1D numerical software package [9]. A sensitivity analysis is carried out of various
factors affecting radionuclide transport, and indirectly the risk
assessment, including soil texture and especially preferential ﬂow

E.M. Pontedeiro et al. / Journal of Hazardous Materials 174 (2010) 648–655

through the waste layer, the vadose zone and the granular aquifer
below and downgradient from the site.
2. Mining site and disposal facility
The radionuclide transport and safety assessment is applied
here to a still operating NORM waste site located in the northof Brazil close to the equator. The analysis permits an evaluationof the potential consequences of slag disposal in a tropical forest. The exact location of the site is not given here for reasons of
conﬁdentiality. Ore extracted from an open pit mine at the site contains natural radionuclide decay chains, with the daughter products
mostly in secular equilibrium with the parent. Different physicaland chemical procedures (crushing/grinding, separation, concentration) are employed at the site to produce a ﬁnal concentrate forpyrometallurgical processing. Pyrometallurgy involves treatment
in a furnace at high temperatures to separate metals from large
amounts of waste rock still present in the concentrate. The waste
is most often removed in the form of slags in which all radioactive
species have become more concentrated as compared to the original ore, except for radioactive lead which, together with its stable
isotope, may be volatilized during the production process.
The radioactive components of the slag have long decay chains(14 daughters for the 238 U series) that end in stable lead. Modelingthe transport of these contaminants hence should involve not only
the parent nuclide, but also all of its daughter products. The calculations with the decay chains are not trivial since the daughter
radionuclides formed during the decay process all have different
physical and chemical properties. The daughters hence may movefaster or slower than the parent in different layers of the disposal
system and groundwater, while also having radiological toxicities
that differ from those of the parent.
The concentrations of short-lived daughters were calculated
from the activity of the parent nuclides (directly measured) using
the assumption of secular equilibrium and applied to elements withhalf-lives of not more than 1 year. The general parent/daughter
decay chain is of the form:
238

U → 234 Th → 234 Pa → 234 U → 230 Th → 226 Ra → 222 Rn→ 218 Po → 214 Pb → 214 Bi → 214 Po → 210 Pb → 210 Bi→ 210 Po → 206 Pb(stableelement)
In our analysis we limited ourselves to the chain 238 U → 234 U→ 230 Th → 226 Ra → 210 Pb. Activities of the various isotopes arereported here in units of Bq/l for liquid-phase concentrations andBq/g for solid-phase concentrations.
The facility containing the slags consisted of a modular disposal
unit constructed of earth materials. An impermeable liner placed
under the waste layer (a requirement of the regulatory agency)
was not considered in our analysis. Average annual precipitation
measured at a location close to site was 2430 mm, and the average
evapotranspiration rate (as calculated with the Penman–Monteithequation) was 1610 mm. Adjusted for runoff, the long-term average recharge rate at the site was estimated to be 657 mm/y. This
recharge rate was used as a surface ﬂux boundary condition for thesteady-state simulations of water ﬂow through the site and into
groundwater.
The dimensions of the disposal unit were 70 m wide by 100 mlong and 6.0 m deep. Initial radionuclide concentrations were
71 Bq/g for 238 U and 234 U, 67 Bq/g for 230 Th, 63 Bq/g for 226 Ra and
4.8 Bq/g for 210 Pb. These values suggest that U is likely in secularequilibrium with 230 Th. Since 226 Ra is more mobile than the otherradionuclides, there will be a loss of Ra by leaching. Because also
some 222 Rn will escape through volatilization, both 226 Ra and 222 Rn
will not be in equilibrium with 210 Pb during the pyrometallurgi-649

cal process, and radioactive lead will volatize simultaneously withstable lead.
The slag waste layer had a measured bulk density of 1.89 g/cm3 .
The unsaturated hydraulic properties of the waste layer and theunderlying vadose zone were described using the constitutive
equations [10]:Se (h) =(h) − r1=n m
s − r[1 + (˛h) ]1/m m

K(h) = Ks Se0.5 [1 − (1 − Se

) ]



m=1−

1
n


(1)

2

(2)

where Se is effective saturation, h is the pressure head,  is thevolumetric water content,  r and  s represent the residual and saturated water contents, respectively, Ks is the saturated hydraulicconductivity, and ˛ and n are empirical shape parameters [10].
Values of the hydraulic parameters of the waste layer were estimated from long-term drainage experiments on large lysimeters
containing slags from the site, operated by the Poços de Caldas
Laboratory of the Brazilian Nuclear Energy Commission (CNEN) in
Minas Gerais, Brazil.
We also tried to estimate distribution coefﬁcients (Kd values) of
the radionuclides from the lysimeter experiments with the slags,
but the concentrations (except those of 226 Ra) of the leachate were
too low and too variable to obtain reliable estimates. Preliminary
laboratory batch measurements at the Poços de Caldas Laboratory
indicated radionuclide Kd values very close to those of a clay soil[11], except for Ra, which was closer to that of sand. For these reasons we used Kd values typical of clay for U, Th, and Pb (1.6, 5.8, and0.54 m3 /kg, respectively), and the measured value (0.88 m3 /kg) for
Ra. The same values were initially used for all layers (slag, vadosezone, aquifer). The assumed Kd values produced very high retardation factors (R) factors for all of the radionuclides involved (up to
about 40,000 for Th), and hence very low effective transport rates.
In another analysis we also used the hydraulic parameters and Kdvalues of a typical coarse-textured (sandy) soil to show the verysigniﬁcant effects of soil texture on the simulations.
The unsaturated zone below the waste layer consisted of reddish
Belterra clay as described by Dennen and Norton [12] and Truckenbrodt and Kotschoubey [13]. Although Oxisols such as thosefound at the site can have very high clay contents (up to 90%),
they often show relatively high inﬁltration rates more typical ofcoarse-textured soils, and considerable macroporosity. Soil texture(clay), bulk density (1.3 g/cm3 ) and the saturated hydraulic conductivity (Ks = 21 m/y) of the vadose zone were locally measured. Other
hydraulic parameters in Eq. (1) for Belterra clay were taken from
Belk et al. [14], except for the value of ˛ (we used 4.5 m−1 ), which
was incorrectly printed in [14]. The hydraulic parameters in the
Belk et al. paper were otherwise very much consistent with genericvalues suggested by Hodnett and Tomasella [15] and Tomasella etal. [16].
The saturated zone consisted of red saprolite resulting fromin-situ alterations of acid volcanic rocks, interspersed with lightercolored (white, yellow) lenses derived from feldspars. The phreaticaquifer had an average thickness of 4.5 m, with a natural hydraulicgradient of 0.05625. Geotechnical essays of the aquifer produced
values for the saturated hydraulic conductivity (Ks ) between
4.22 × 10−5 and 7.9 × 10−4 cm/s. For our calculations we used avalue of 1.7 × 10−4 cm/s for clay soil textural class as estimated withthe pedotransfer function module of HYDRUS-1D, which is basedon the hierarchical neural network (Rosetta) approach of Schaap etal. [17].
For the calculations we assumed a 6-m thick homogeneous
waste layer (overlain by a negligibly thin layer of local soil), and a 5m thick unsaturated zone consisting of Belterra clay. Radionuclides
leached vertically from the vadose zone were assumed to mix with

650

E.M. Pontedeiro et al. / Journal of Hazardous Materials 174 (2010) 648–655

laterally moving groundwater. The mixing process was describedusing the EPA mixing zone model [18]. Radiological doses and risks
were calculated using a program for the biosphere based on recommendations by ISAM [11] and programmed [3] using the symbolic
computational software Mathematica 5.2 [19]. We did not impose
any time limits (e.g., 50,000 years) on our simulations; calculations
for each scenario and radionuclide of interest in some cases continued until the peak value of the relevant safety indicator (dose)
had been reached.
3. Simulation methodology
Water ﬂow and radionuclide transport through the waste andvadose zone were simulated using the HYDRUS-1D software package [9]. Simulations assumed applicability of the Richards equation
for ﬂow in variably-saturated media as follows:
∂(h)
∂
=
∂t
∂z



K(h)





∂h
−1
∂z

(3)

where z is the vertical coordinate positive downwards, t is time, h is
the pressure head,  is the volumetric water content, and K(h) is theunsaturated hydraulic conductivity function given by Eq. (2). The
transport of each radionuclide within the decay chain was initiallymodeled using the equilibrium advection–dispersion equation:∂(RC)
∂
∂C
∂qC
=
( D
)−
+
∂t
∂z
∂z
∂z

(4)

where C is the solution concentration, R (=1 + Kd /) is the solute
retardation factor,  is the material bulk density, Kd is the distribution coefﬁcient for linear equilibrium sorption, D is the solute
dispersion coefﬁcient, and q is the volumetric ﬂuid ﬂux density.
The general reaction term  represents ﬁrst-order source-sink coupling of two consecutive radionuclides within the decay chain, withdecay occurring in both the solution and adsorbed phases.
Eq. (4) assumes equilibrium transport. Unfortunately, much evidences exist that ﬂuid ﬂow and contaminant transport processes inthe subsurface generally cannot be described using classical models that assume equilibrium or uniform ﬂow conditions [20,21]. A
large number of dual-porosity and dual-permeability models havebeen proposed to simulate preferential and/or non-equilibrium
ﬂow in soils and groundwater [21–23]. Here we limit our analysis
to the concept of two-region (dual-porosity) transport involving
physical non-equilibrium [24,25]. This approach assumes that theliquid-phase can be partitioned into mobile (macroporous,  m ), andimmobile (microporous or matrix,  im ) regions, with  =  m +  im .
Advective–dispersive transport is allowed only in the mobile phase,
while the transfer of solute between the mobile and immobile
region is controlled by diffusion. Assuming mobile/immobile typephysical non-equilibrium, the governing transport equations foreach radionuclide in the decay chain are now of the form [25]∂(m + fKd )Cm
∂
=
∂t
∂z



m Dm

∂Cm
− qCm
∂z



− ˇ(Cm − Cim ) + m
(5)∂[im + (1 − f )Kd ]Cim
= ˇ(Cm − Cim ) + im∂t(6)where the subscripts m and im refer to the mobile and immobile
liquid phases, respectively, f denotes the fraction of sorption sites
in contact with mobile water, and ˇ is the average mass transfer
coefﬁcient for solute exchange between the mobile and immobile
phases.
Eqs. (5) and (6) allow simulation of preferential or nonequilibrium transport in the subsurface with a minimum of

Table 1
Estimated values of the mass transfer coefﬁcient, ˇ, for radionuclide transport
through the NORM waste and the Belterra clay subsurface.
238−1

ˇwaste (y )ˇvadose (y−1 )ˇgroundwater (y−1 )U/234 U

0.00160.00250.0008230Th0.00060.00090.0003226

Ra

0.00260.00410.0013210

Pb

0.00390.00600.0019additional transport parameters. The model contains three additional parameters relative to the equilibrium model:  m (or  im ),f, and ˇ. Following Nkedi-Kizza et al. [26], f was set equal to therelative fraction of mobile water (i.e., f =  m /), thus reducing thenumber of additional parameters to 2 ( m / and ˇ). In all of thecalculations reported here we used a value of 0.5 for the fractionof mobile water,  m /, which contrasts well with the equilibrium
approach ( m / = 1). For ˇ we performed a sensitivity analyses using
values from 0.001 (far from equilibrium) to 1 (close to equilibrium).
In our ﬁnal analysis we used values recommended by Maraqa [27]based on an analysis of previously published laboratory and ﬁeldscale transport experiments. Maraqa found that the mass transfer
coefﬁcient ˇ is inversely related to the residence time of the contaminant in the system as follows:ˇ = atr−b(7)where tr is the residence time in hours, and b is in the range
0.73–0.88 (we assumed a value of 0.8). Maraqa [27] estimated thevalue of a to be 0.95 for ﬁne non-aggregated media (assumed in
our analysis), and 0.26 for aggregated media. Estimated values of ˇfor each radionuclide are listed in Table 1. Residence times in eachof the three subsystems (waste layer, vadose zone, groundwater)
were estimated using tr = LR/v, where L is the length of the subsystem, R is the retardation factor, and v is the average pore watervelocity (v = q/). Dispersivities were assumed to be one tenth of thelength of each subdomain, while radionuclide half-life data weretaken from ICRP [28].
4. Transport simulation results
The performance assessment of the disposal facility was carriedout using a leaching and off-site small farm scenario. The system
was modeled by assuming that rainfall percolated vertically downward through the disposal site and the unsaturated zone, and thenmixed with and moved laterally in the aquifer. Radionuclides transported in groundwater were assumed to be intercepted by a well,located 100 m downgradient of the landﬁll. For the safety analysis
we only used concentrations of the aquifer at the well point.
The ﬁrst case considered equilibrium transport in each of thethree media (the waste layer, the vadose zone, and groundwater).
Calculated concentrations are shown in Fig. 1. Notice that 238 U and
234 U give by far the highest concentrations at both the bottom of
the vadose zone and at the well. However, some contribution of
230 Th is also clearly present in the leachate from the vadose zone,
but not at the well.
In order to analyze the impact of preferential ﬂow on the ﬁnal
safety assessment, different cases were simulated using a mobile
fraction of water (and an f value) of 0.5 separately in each of thethree layers. Figs. 2–4 show the effects of preferential ﬂow in,
respectively, the waste layer, the vadose zone only, and the aquifer
only, assuming different values of the mass transfer coefﬁcient ˇ.
The highest concentrations (especially for U and Th) in the leachateleaving the waste layer were obtained when considering equilibrium (Fig. 2), or with relatively high values of the mass transfer
coefﬁcient ˇ (i.e., situations near equilibrium). This shows that preferential ﬂow (small values of ˇ) of rain water through the waste
leads to less rapid pollution of the subsurface (i.e., rain water par-

E.M. Pontedeiro et al. / Journal of Hazardous Materials 174 (2010) 648–655

651

Fig. 1. Calculated radionuclide concentrations versus time leaving the vadose zone (left) and at the well (right) assuming equilibrium transport (Belterra clay subsurface).

tially bypasses the waste). As compared to U and Th, Ra and Pb wereboth found to decay very quickly.
Fig. 3 shows the effects of preferential ﬂow through the unsaturated zone, while assuming equilibrium transport in the slags layer.
Concentrations at the bottom of the vadose zone now progressively
increase when ˇ decreases (more preferential ﬂow). Notice that thefront ends of the plumes arrive earlier for U and Th, while for Ra andPb the concentration peaks are higher because of shorter half-lives.
The effects of preferential ﬂow in the aquifer (but with equilibrium transport in the slags and unsaturated zone) are shown in
Fig. 4. Concentrations here pertain to a well 100 m downgradientfrom the waste site. The ﬁve radionuclides all behave very similarly, with the most rapid transport and highest concentrations
occurring when ˇ = 0.001 y−1 . The ﬁrst peaks for 226 Ra and 210 Pb atabout 15,000 years were mostly due to advective–dispersive transport, while the increases after 30,000 years were a consequence oftransport and decay of the predecessor radionuclides in the decay
chain.
The sensitivity analysis for ˇ shows interesting results in that
preferential ﬂow of rain water through the waste layer leads tolower concentrations in the unsaturated zone below the waste
layer, and in groundwater. On the other hand, preferential ﬂow

Fig. 2. Sensitivity analysis for preferential ﬂow in the waste layer (Belterra clay).

652

E.M. Pontedeiro et al. / Journal of Hazardous Materials 174 (2010) 648–655

Fig. 3. Sensitivity analysis for preferential ﬂow through the unsaturated zone below the waste layer (Belterra clay).

in the unsaturated and groundwater regions causes more rapidtransport of the radionuclides, and higher peak concentrations.
Fig. 5 shows the totality of the effects of preferential ﬂow on the
results using the values of ˇ as predicted with Eq. (7) and listed

in Table 1. A comparison with the results for equilibrium transport(Fig. 1) shows that the peak concentrations for uranium (having thehighest concentrations) are only marginally affected. However, thepeak concentrations at the bottom of the vadose zone and espe-

Fig. 4. Sensitivity analysis for preferential ﬂow in the phreatic aquifer towards the well (Belterra clay).

E.M. Pontedeiro et al. / Journal of Hazardous Materials 174 (2010) 648–655

653

Fig. 5. Calculated radionuclide concentrations leaving the vadose zone (left) and at the well (right) for preferential ﬂow using values of ˇ listed in Table 1 (Belterra clay).

Fig. 6. Calculated radionuclide concentrations leaving the vadose zone (left) and at the well (right) assuming equilibrium transport (sand subsurface).

cially at the well (after 43,000 years for preferential ﬂow versus
83,000 years for equilibrium transport) occur now much earlier.
The radionuclides also arrived much earlier at the well (12,500
years for preferential ﬂow versus 22,000 years for equilibriumtransport).
All of the simulations so far were carried out for Belterra clay,
leading to very conservative values for the hydraulic conductivity(permeability) and especially the estimated Kd values, and hence
the contaminant transport rates. A complementary set of simulations was carried out to show the impact of soil texture on thelong-term concentrations. For these simulations we assumed thepresence of a very coarse-textured subsurface below the waste
layer (i.e., for the vadose zone and groundwater), while leaving
the physical and chemical properties of the waste layer itself unaltered. For Kd we selected typical literature values for sand: 0.035,
3.2, 0.5 and 0.27 m3 /kg for U, Th, Ra and Pb, respectively, taken from
ISAM [11]. For the unsaturated hydraulic properties in Eqs. (1) and(2) we used typical values for sand using the Rosetta module of
HYDRUS-1D.
Results for equilibrium transport from the waste layer throughthe now sandy subsurface (vadose zone and aquifer) are shown in
Fig. 6. As compared to the simulations for Belterra clay in Fig. 1,
the 238 U/234 U peaks reach the bottom of the vadose zone muchearlier (17,000 versus 24,000 years), while the peak concentration
itself is 5.95 × 106 Bq/m3 , more than 100 times that for the clay soil
(3.72 × 104 Bq/m3 ). This reﬂects the much faster transport throughthe sandy vadose zone. Lateral transport rates through the aquifer
are similarly much faster, with the 238 U/234 U peaks reaching thewell 100 m downstream of the site now after only 18,000 years
with a peak concentration of 4.92 × 106 Bq/m3 , compared to 85,000
years and a peak concentration of 1.4 × 104 Bq/m3 for the Belterra
clay subsurface.
We next superimposed the effects of preferential ﬂow on theequilibrium results for the sandy subsurface shown in Fig. 6. Asbefore, we estimated the values of the residence times in the vadosezone and the aquifer using Eq. (7), and then estimated the masstransfer coefﬁcients using the correlations derived by Maraqa [8]for non-aggregated media (shown in Table 2). The value of f was set
at 0.5, the same as for the clay soil. The results for preferential ﬂow
are shown in Fig. 7. Comparing these results with those in Fig. 6
for equilibrium transport shows that, similarly as for the Belterra
clay subsurface, the peak concentrations are again only marginallyhigher for the preferential ﬂow scenario (e.g., 4.6 × 106 Bq/m3 forthe sandy vadose zone compared to 3.7 × 104 Bq/m3 for the clay).
The peak concentrations however occur much earlier, especially at
the well (11,000 years for the sandy subsurface versus 43,000 years
for the clay). This shows that both soil texture and preferential ﬂow
can have major impacts on the radionuclide transport rates.

Table 2Estimated values of the mass transfer coefﬁcient, ˇ, for radionuclide transport
through the sandy subsurface of the waste disposal problem.
238−1

ˇvadose (y )ˇgroundwater (y−1 )U/234 U

0.054
0.34

230

Th

0.00150.0093226

Ra

0.0065
0.041

210

Pb

0.011
0.067

654

E.M. Pontedeiro et al. / Journal of Hazardous Materials 174 (2010) 648–655

Fig. 7. Calculated radionuclide concentrations leaving the vadose zone (left) and at the well (right) for preferential ﬂow (sand subsurface).

5. Safety assessment
The primary objective of a performance assessment is to provide
evidence that human health and the environment are protected asmuch as possible in the future. This can be done only by carrying out
simulations and using the results to judge if the adopted disposal
concepts meet the radiological standards for long-term protectionof the public and the environment, as established by regulatory
bodies. A pathways analysis with different scenarios provides a systematic way of evaluating potential routes by which the humanpopulation could be exposed to radiation [2,3,8]. Because of theunpredictable nature of the long-term integrity of disposal systems, as well as of human behavior, the post-closure scenarios areinevitably hypothetical.
Our risk assessment of the NORM waste site assumes that thedose projection for members of the public cannot exceed 1 mSv/y.
Assuming a risk factor of 5% per Sievert (Sv) as recommended by
the International Commission on Radiological Protection [28], thisleads to an assessed radiological risk of fatality from the facility
to a representative member of the potentially exposed group of
5 × 10−5 (i.e., 5 in one hundred thousand per year). The acceptablefraction, termed the dose or risk constraint, shall be determinedby the regulatory agency. The Brazilian regulatory body adopted adose constraint of 0.3 mSv/y above natural background.
Our small farm scenario was modeled assuming that radionuclides migrated vertically through the disposal and vadose zones,
and then laterally through the aquifer until intercepted by a wellor discharged into a stream. We assumed that the well or stream
provided the only available source of water for residents for direct
consumption, irrigation or for use by cattle. The radiological dose
was assessed for an exposed critical group of individuals in the population receiving the highest dose or risk. Data regarding humanbehaviour were based on the dietary habits of adults from thatgroup (e.g., water and food consumption, breathing rate, and hours
indoor or outdoor). Residents received doses from external radiation, ingestion of contaminated water and food, and inhalation

Fig. 8. Comparison of annual risk assuming either equilibrium transport or preferential ﬂow (Belterra clay).

E.M. Pontedeiro et al. / Journal of Hazardous Materials 174 (2010) 648–655

of airborne radionuclides transported from the waste site or suspended in runoff water following irrigation. Following Pontedeiro
et al. [3], the following speciﬁc processes in the biosphere wereconsidered: (a) ingestion of water from the well, (b) irrigation, (c)re-suspension or particulates and inhalation, (d) external radiation
exposure, (e) consumption of home-grown produce, (f) consumption of contaminated meat, (g) ingestion of contaminated milk, (h)accidental ingestion of contaminated soil, (i) inhalation of radonand decay products from soil, and (j) contact with surface water,
including transfer to ﬁsh and humans. Ingestion, inhalation andexternal dose conversion factors (Sv Bq−1 ), as well as transfer coefﬁcients and concentration factors were taken from [11].
The risk assessment was performed for both equilibriumtransport and the situation where preferential ﬂow occurs simultaneously in all three subsurface systems (waste, vadose zone,
groundwater). Results for the Belterra clay subsurface are presented in Fig. 8. The calculations show that Uranium is mostly
responsible for the total dose up to 80,000 years. After that time
the contribution of 230 Th becomes increasingly more signiﬁcant,leading to approximately the same levels of total risk. The resultsin Fig. 8 further show that preferential ﬂow causes the risks to shift
to much earlier times, but that the maximum of the total annual
risk is not substantially affected. This is also true for uranium but
not for thorium and radium, which both show increasing annual
risks over the considered time period of 100,000 y. Similar conclusions about the shifts to earlier times and the peak concentrations
also hold if the subsurface was made up of more coarse-textured
material, but such results are not shown here.
6. Conclusions
A major objective of this study was to illustrate the importance of having both a comprehensive process-based descriptionof the fate and transport of NORM decay chains in the subsurface,and a detailed risk or dose analysis over relatively long periodsof time (e.g., 50,000–100,000 years). The reliability of quantitative
safety indicators such as those used here should of course decrease
with time due to inherent uncertainties about future environmental conditions and human activities. The maximum annual risk overthe 100,000 years period was approximately 5 × 10−5 (a dose of
1 mSv/y) for the Belterra clay subsurface at the Amazonian NORM
waste site considered in this study. Since the dose constraint for
NORM installations in Brazil is 0.3 mSv/y above natural background,the maximum calculated total doses and associated risks in ourcase study were always higher, irrespective of the assumption of
having equilibrium or preferential ﬂow in the subsurface, whena time frame of 100,000 years is considered. However, the maximum annual dose limit was not exceeded when restricting the
calculations to only the ﬁrst 50,000 years and assuming equilibriumtransport.
Risk assessments are usually carried out assuming equilibriumtransport in the subsurface. In this study we showed that preferential ﬂow and soil texture can have a major effect by accelerating thetransport of radionuclides in the unsaturated zone and the aquifer,thus affecting resultant concentrations and dose calculations. Similar major effects on transport and risk analyses are to be expected
when other non-ideal transport processes are considered, such as
chemical non-equilibrium and colloid-facilitated transport.
References[1] C.J. Roberts, J.B. Quinby, W.P. Duggan, Y. Yuan, Disposal options and case-study
pathway analyses, Appl. Radiat. Isotopes 49 (3) (1998) 241–258.

655[2] Extent of environmental contamination by naturally occurring radioactive
material (NORM) and technological options for mitigation, Technical Report
Series No. 419, Int. Atomic Energy Agency, IAEA, Vienna, 2003 (available at
http://www-pub.iaea.org/MTCD/publications/Pubdetails.asp?pubId=6789).
[3] E.M. Pontedeiro, P.F.L. Heilbron, R.M. Cotta, Assessment of mineral industry
NORM/TENORM disposal in hazardous landﬁlls, J. Hazard. Mater. B139 (2007)
563–568.
[4] E.R. Landa, Naturally occurring radionuclides from industrial sources: characteristics and fate in the environment, Radioactivity Environ. 10 (2007) 211–237(Elsevier).
[5] P.R. Schroeder, T.S. Dozier, P.A. Zappi, B.M. McEnroe, S.J. Sjostrom, R.L.
Peyton, Hydrologic Evaluation of Landﬁll Performance (HELP) Model,
EPA/600/R-94/168, U.S. Environmental Protection Agency, Washington, DC,
1994.
[6] K. Pruess, The TOUGH codes; a family of simulation tools for multiphase ﬂow
and transport processes in permeable media, Vadose Zone J. 3 (2004) 738–746.
[7] J. Simunek, M.Th. van Genuchten, M. Sejna, Development and applications of
the HYDRUS and STANMOD software packages and related codes, Vadose Zone
J. 7 (2008) 587–600.
[8] C. Yu, A.J. Zielen, J.-J. Cheng, Y.C. Yuan, L.G. Jones, D.J. LePoire, Y.Y. Wang, C.O.
Loureiro, E. Gnanapragasam, E. Faillance, A. Wallo III, W.A. Williams, H. Peterson, Manual for implementing residual radioactive material guidelines using
RESRAD, Version 5.0. ANL/EAD/LD-2, Argonne National Laboratory, Argonne,
IL, 1993.
[9] J. Simunek, M.Th. van Genuchten, M. Sejna, The HYDRUS-1D Software Package
for Simulating the One-Dimensional Movement of Water, Heat and Multiple Solutes in Variably-Saturated Media, Version 3.0. HYDRUS Software Series
1, Department of Environmental Sciences, University of California, Riverside,
2005, p. 240.
[10] M.Th. van Genuchten, A closed-form equation for predicting the hydraulic conductivity of unsaturated soils, Soil Sci. Soc. Am. J. 44 (1980) 892–898.
[11] Derivation of quantitative acceptance criteria for disposal of radioactive waste
to near surface facilities: development and implementation of an approach,
ISAM Safety Report, Version 3.0, IAEA, Vienna, Austria, March 1999.
[12] W.D. Dennen, H.A. Norton, Geology and geochemistry of bauxite deposits inthe lower Amazon basin, Econ. Geol. 72 (1977) 82–89.
[13] W. Truckenbrodt, B. Kotschoubey, Cobertura terciária das bauxitas amazônicas,
Revista Braileira Geociências 11 (1981) 203–208.
[14] E.L. Belk, D. Markewitz, T.C. Rasmussen, E.J.M. Carvalho, D.C. Nepstad, E.A.
Davidson, Modeling the effects of throughfall reduction on soil water content
in a Brazilian Oxisol under a moist tropical forest, Water Resour. Res. 43 (2007),
doi:10.1029/2006WR005493 (W08432).
[15] M.G. Hodnett, J. Tomasella, Marked differences between van Genuchten soil
water retention parameters for temperate and tropical soils: a new waterretention pedo-transfer function developed for tropical soils, Geoderma 108(2002) 155–180.
[16] J. Tomasella, Ya. Pachepsky, S. Crestana, W.J. Rawls, Comparison of two techniques to develop pedotransfer functions for water retention, Soil Sci. Soc. Am.
J. 67 (2003) 1085–1092.
[17] M.G. Schaap, F.J. Leij, M.Th. van Genuchten, Neural network analysis for hierarchical prediction of soil hydraulic properties, Soil Sci. Soc. Am. J. 62 (1998)
847–855.
[18] Soil Screening Guidance: Technical Background Document, EPA/540/R-95/128,
PB96-963502, Ofﬁce of Solid Waste and Emergency Response, U.S. Environmental Protection Agency, Washington, DC, 1996.
[19] S. Wolfram, Mathematica 5.2, Wolfram Media, Cambridge, UK, 2005.
[20] J.M.H. Hendrickx, M. Flury, Uniform and preferential ﬂow mechanisms inthe vadose zone, in: Conceptual Models of Flow and Transport in the Fractured Vadose Zone, National Academy Press, Washington, DC, 2001, pp. 149–
187.
[21] N.J. Jarvis, Review of non-equilibrium water ﬂow and solute transport in soil
macropores: principles, controlling factors and consequences for water quality,
Eur. J. Soil Sci. 58 (2007) 523–546.
[22] J. Simunek, M.Th. van Genuchten, Modeling nonequilibrium ﬂow and transport
processes using HYDRUS, Vadose Zone J. 7 (2008) 782–797.
[23] J.M. Köhne, S. Köhne, J. Simunek, A review of model applications for structured
soils: a) Water ﬂow and tracer transport, J. Contam. Hydrol. 104 (2009) 4–
35.
[24] M.Th. van Genuchten, P.J. Wierenga, Mass transfer studies in sorbing porous
media. I. Analytical solutions, Soil Sci. Soc. Am. J. 40 (1976) 473–480.
[25] M.Th. van Genuchten, R.J. Wagenet, Two-site/two-region models for pesticidetransport and degradation: Theoretical development and analytical solutions,
Soil Sci. Soc. Am. J. 53 (1989) 1303–1310.
[26] P. Nkedi-Kizza, J.W. Biggar, M.Th. van Genuchten, P.J. Wierenga, H.M. Selim, J.M.
Davidson, D.R. Nielsen, Modeling tritium and chloride 36 transport through anaggregated oxisol, Water Resour. Res. 19 (1983) 691–700.
[27] M.A. Maraqa, Prediction of mass-transfer coefﬁcient for solute transport inporous media, J. Contam. Hydrol. 53 (2001) 153–171.
[28] Radionuclide Transformations Energy and Intensity of Emissions, Int. Commission on Radiological Protection ICRP Publication 38, Pergamon Press, Oxford,
1983.