The U.S. Environmental Protection Agency, through its Office of Research and Development,
funded and managed the research described here under Contract #EP-D-11-006 to Eastern
Research Group. It has been subjected to the Agency's review and has been approved for
publication. Note that approval does not signify that the contents necessarily reflect the views of
the Agency. Mention of trade names, products, or services does not convey official EPA
approval, endorsement, or recommendation.

Questions  concerning this document or its application should be addressed to:
Sang Don  Lee, Ph.D.
National Homeland Security Research Center
Office of Research and Development
United States Environmental Protection Agency
109 T W Alexander Drive
Durham, NC 27705


AGRICP     Agricultural Countermeasure Program
Am          Americium
AMAD       Activity Median Aerodynamic Diameter
Ba           Barium
CBRN       Chemical, biological, radiological, and nuclear
Ce           Cerium
Cf           Californium
Co           Cobalt
CONDO     CONsequences of Decontamination Options
Cs           Cesium
DIPCOT     DIsPersion over COmplex Terrain
EA          Environment agency
ERMIN      European Model for Inhabited Areas
Eu           Europium
FSA         Food Standards Agency
GIS          Geographic information system
I            Iodine
Ir            Iridium
IS           Information system
JAERI       Japan Atomic Energy Research Institute
MAA        Multi-Attribute value Analysis
Mo          Molybdenum
MOGRA     Migration Of GRound Additions
MOIRA      The MOdel based computerized system for management support to Identify
             optimal remedial strategies for Restoring radionuclide contaminated Aquatic
NPP         Nuclear power plant
NW          Nuclear weapon
PARATI     Program for the Assessment of RAdiological consequences in a Town and of
             Intervention after a radioactive contamination
Po           Polonium
Pu           Plutonium
Ra           Radium
ROD         Radiological dispersal device
RESRAD     RESidual RADioactive
RODOS      The Real-time On-line DecisiOn Support
Ru           Ruthenium
Sr           Strontium
Te           Tellurium
Zr           Zirconium

Disclaimer	ii
Abbreviations and Acronyms	iii
Table of Contents	iv
Figures and Tables	vi
Acknowledgements	vii
Executive Summary	1
1.   Introduction	2
  1.1.  Methodology	2
2.   Sources of Radiation	3
  2.1.  Nuclear Weapons	4
  2.2.  Nuclear Power Plants	5
  2.3.  Radiological Dispersion Devices	6
3.   Spreading Mechanisms	7
  3.1.  Deposition	7
  3.2.  Surface Transport	8
    3.2.1.   Urban Environments	8
    3.2.2.   Buildings	9
    3.2.3.   Streets and Sidewalks	9
    3.2.4.   Sewers	9
    3.2.5.   Parks	10
  3.3.  Rural Environments	10
    3.3.1.   Forests	10
    3.3.2.   Soil	11
    3.3.3.   Aquatic Environments	12
4.   Resuspension	13
  4.1.  Natural Resuspension Causing Mechanisms	14
    4.1.1.   Wind	14
    4.1.2.   Rain	15
    4.1.3.   Fire	15
  4.2.  Human Resuspension-Causing Mechanisms	15
    4.2.1.   Vehicles	16

    4.2.2.   Pedestrians	17
    4.2.3.   Agricultural Activities	17
    4.2.4.   Decontamination Activities	17
  4.3.   Factors Influencing Resuspension	17
    4.3.1.   Particle Size	17
    4.3.2.   Range	17
    4.3.3.   Timeline	18
    4.3.4.   Limiting Factors	18
    4.3.5.   Climates	18
    4.3.6.   Seasons	18
5.   Modeling and Simulation	19
  5.1.   ARGOS	19
  5.2.   ERMIN	20
  5.3.   RODOS	21
  5.4.   CONDO	22
  5.5.   MOIRA	22
  5.6.   RESRAD	23
  5.7.   MOGRA	23
  5.8.   PARATI	23
6.   Discussion	24
  6.1.   Source Characterization	24
  6.2.   Surface Migration	24
  6.3.   Resuspension	25
  6.4.   Modeling and Simulation	26
  6.5.   Conclusion	26
    6.5.1.   Spreading Mechanisms	27
    6.5.2.   Modeling and Simulation	28
7.   References	29


Figure 1. Resuspension Mechanisms According to Natural and Human Activities	14

Table 1. Particle Characteristics According to Fallout Type [7]	5

Table 2. Mean Depth Profiles for 240Pu and 239Pu, 90Sr, and 137Cs [52, 58-60]	12


Contributions of the following individuals and organizations to this report are gratefully

Peer Reviewers:

Scott Hudson, U.S. Environmental Protection Agency, Office of Solid Waste and Emergency
James Mitchell, U.S. Environmental Protection Agency, Region 5
Terry Stilman, U.S. Environmental Protection Agency, Region 4

Key Contributors:

Sang Don Lee, U.S. Environmental Protection Agency, Office of Research and Development
Timothy Boe, Eastern Research Group
Colin Hayes, Eastern Research Group


With the prevalence of terrorism, the threat of an attack targeting densely populated urban areas
is of increasing concern. One possible means of attack is a radiological dispersal device
(RDD)—essentially, a dispersal device (e.g., an explosive) capable of spreading contaminants
over a confined area. RDDs are unlikely to invoke mass casualties; instead, they contaminate
people and surfaces with radioactivity. Although the activity levels of the resulting
contamination may be low, chronic exposure could result in damaging health effects. Unlike
other sources of radioactivity (i.e., nuclear power plant [NPP] accidents and nuclear weapon
[NW] detonations), RDDs are extremely mobile and relatively simple to build. However, very
little is known about the nature of contaminants produced by RDDs or their behavior in urban
environments. In contrast, a considerable amount of research has been conducted on NPP
accidents and NW detonations. This paper seeks to compare NPP and NW incidents and their
derived contaminants to better understand RDDs and how they might interact with urban

This literature review was conducted to address the current state of knowledge on particle
transport relative to radiological sources and their host environments. More specifically, this
review seeks to 1) determine whether empirical evidence exists for further characterizing RDDs
according to literature pertaining to NW detonations and NPP accidents, 2) provide an overview
and analysis of the current state of knowledge related to radiological sources with reference to
particle transport, 3) contrast the behavior of radionuclides in urban and rural environments, and
4) explore the current state of radiological transport models, taking the above findings into
consideration. The literature review concludes by discussing research gaps and research needs in
order to improve response and recovery capabilities following a RDD incident.

Numerous knowledge gaps, as identified by this review, need to be addressed in order to better
understand and predict the transport of radiological contaminants. The gaps include, but are not
limited to, the impacts of natural and human-derived resuspension within urban environments
and the necessary countermeasures to reduce the impacts of these mechanisms, water-soluble
radioactive contaminant transport within urban  areas, surface interaction of dispersed particles,
and a thorough technical review of available radiological models with reference to particle


The release of radiological material from an incident has the potential for widespread
contamination and public exposure. The magnitude of such an event may be augmented by the
continued transport of radioactive material. In a relatively short time, contaminants may migrate
to otherwise clean areas, increasing the extent of contamination. This phenomenon was recently
observed near Fukushima, in which contaminants migrated to uncontaminated or decontaminated
areas [1]. In essence, there are few options for sequestering atmospheric transport or surface
migration. Short-term gross mitigation efforts, let alone long-term cleanup, would  likely require
a significant amount of resources and time.

In addition to the environment in which the radioactive contaminants are released,  the source of
contamination will influence the magnitude and overall behavior of radioactive material.
Radiological dispersal devices (RDDs) are one such source.

RDDs are designed to disperse radioactive material in order to expose people and contaminate
areas. Although their ability to contaminate large areas or expose vast numbers of people to
radiation is limited compared to nuclear power plant (NPP) accidents or nuclear weapon (NWs)
detonations, RDDs are capable of contaminating targeted areas of interest, disrupting normal
activities, inflicting major economic implications, and invoking fear [2]. The radioactive material
needed to construct a ROD can be easily appropriated. Such materials are often found in
industrial, research, and medical wastes [3].

Large-scale tests assessing the dispersion and surface transport of radioactive particles associated
with RDDs are limited due to their hazardous nature. Accordingly, there is very little literature
documenting the probable outcome of such an event, especially regarding urban environments.
There is, however, an abundance of literature addressing historical NW detonations and NPP
accidents. This review assesses that literature,  characterizing NPP accidents and NW detonations
in order to facilitate conjectures about the transport of radionuclides analogous to RDDs in urban
environments. Such data could promote more informed response and recovery decisions when
responding to RDD incidents.

This document provides an overview and analysis of the current state of knowledge related to
radiological sources with reference to radiological material transport in urban environments.
More specifically, it seeks to 1) determine whether empirical evidence exists for further
characterizing RDDs according to literature specific to NW detonations and NPP accidents, 2)
provide an overview and analysis of the current state of literature related to radiological sources
with reference to particle transport, 3) contrast the behaviors of radiological materials in urban
and rural environments, and 4) explore the current state of radiological transport models, taking
the above findings into consideration.

1.1. Methodology

To prepare this document, the authors aggregated literature and other information from the
appropriate sources. These sources were queried using a set of predefined keywords. Each piece

of literature was read, assessed, and documented based on a number of criteria (i.e., applicability,
accuracy, clarity, and uncertainty and variability). Literature deemed at least moderately
relevant, according to the above criteria, were then summarized. This document was synthesized
by incorporating the main points presented by pertinent literature.


The detonation of Trinity, the first atomic bomb, began the Atomic Age [4]. The era promised
revelations in defense and energy, many of which came to bear. Over the past 60 years,
radioactive materials have introduced a wide range of beneficial uses, especially in the areas of
medicine, industry, and research. With this modernization came great power—and responsibility,
because conventional and unconventional radioactive sources can adversely impact human health
and the environment  for decades or longer. Radiological incidents can be propagated through
natural and man-made causes and are derived from an array of sources such as NW detonations,
NPP accidents, transportation incidents, sabotage, improvised nuclear devises, and RDDs. Each
radiological incident  varies in terms of release mechanism, release dimensions, deposition
behaviors, isotopic compositions of source materials, and subsequent impact [5].

Radionuclides released into the air may be in a gaseous, particulate, or multi-phase (i.e., gaseous
and particle) form [6]. In particle form,  radionuclides can be characterized by  particle size,
shape, elemental composition, structure (e.g., crystalline and amorphous phases), and valence
and oxidation conditions [7]. The mobility, environmental behavior, bioavailability,  and
ecological impact are determined by the physiochemical properties of the particles themselves.
As particles coagulate or condense onto atmospheric aerosols, their physiochemical form may
change, thus altering their behavior [8]. In short, radioactive particles are distinguished by their
source and formation mechanisms [8]. These characteristics will later influence particle
dispersion, deposition, migration, and remobilization.

For the purpose of this paper, radiological sources are limited to NWs, NPPs,  and RDDs. The
following variables have the greatest influence on the physiochemical characteristics, size, and
structure of particles  in regard to the source:

    •   NWs: device  yield, detonation environment, and height of detonation.
    •   NPPs: oxidative stress, fuel matrix, and release characteristics.
    •   RDDs: elemental and physiochemical properties of radioactive sources, packaging
       material/method, and explosive yield [9].
There are stark differences between the radiological sources listed above. Disparities are most
apparent in incident magnitude, with impacts ranging from small (e.g., RDDs) to large (e.g.,
NWs). The area of contamination increases with incident magnitude. Regardless of
meteorological conditions, the height of detonation and the distance radioactive material is
ejected into the atmosphere will be the main factors that decide the extent of contamination. For
instance, NW detonations can insert particles into the stratosphere by means of extreme
explosive force, where they may traverse the globe in a number of days; particulate originating
from an NPP accident will predominantly reside within the troposphere. Other disparities include

available fuel matrix, activity, and physicochemical characteristics of particles. More
information on this topic is provided in the following sections.

Similarities also exist, especially between NPP accidents and NW detonations. According to
studies that took place near NPPs and NWs, ignoring the destructive nature and extent of
contamination of NW detonations, concentration and deposition values did not vary by more
than one order of magnitude between the two [10]. Similar resuspension values have also been
reported [11]. Furthermore, fallout from NW detonations and NPP accidents is more likely to
produce homogeneous boundaries of contamination, while RDDs are more likely to generate
heterogeneous boundaries[9]. This is based on the assumption that NW detonations and NPP
accidents are more likely to deposit contaminants in a uniform, yet violent manner. The
compositions (i.e., fuel) of NWs and NPPs are somewhat consistent, and therefore are more
likely to produce analogous particle sizes and physiochemical properties. For instance, deposited
contaminants from Fukushima are reportedly similar in isotopic makeup to those from Chernobyl
[12]. As such, these types of incidents are more predictable. On the other hand, RDDs are more
likely to consist of a heterogeneous mixture of fuel and construction material that, when
detonated, will produce a mixture of fragments and particles of variable size and shape leading to
heterogeneous deposition [13].

2.1. Nuclear Weapons

An NW can be described as an explosive device capable of releasing a vast amount of energy in
a limited time and space by means of nuclear interactions [4].  This sudden release of energy
causes extreme increases in pressure and temperature, in turn causing the rapid expanse of
compressed gases (i.e., a shock wave) in addition to the emission of initial, thermal, and residual
radiation [4]. Consequences of NWs may differ greatly in terms of detonation mechanism (i.e.,
fusion vs. fission), composition, and design.

Following detonation of an NW, a large amount of energy is released within a short period of
time. Enveloped weapon residues, nearby aerosols, and ground material are heated to extremely
high temperatures (i.e., ~5,000°C), creating a luminous mass of gaseous material (i.e., a fireball)
[4]. The gaseous particles are eventually solidified by means of nucleation and condensation [4].
As the fireball expands, it engulfs secondary materials (e.g., nearby aerosols and debris).  Once
the contents of the fireball  cool, these secondary materials can serve as nuclei for condensation
of smaller molecules and gas particles [4]. Accordingly, surface area availability is a major factor
in determining particle size distribution, which in turn is influenced by detonation altitude,
device yield, and the surrounding environment.  For instance, ground detonations can produce
very large particle sizes (i.e., up to a few cm in diameter), whereas high-altitude detonations
produce much smaller particle size ranges (i.e.,  approximately 0.01-20 um) [4].

Literature describes four common types of radioactive particles that are characteristic of NW
detonations: 1) condensation of vapors of fission and activation products producing particles
about 1  um in diameter; 2) precipitation of the molten components of the bomb and materials
absorbed by the fireball (e.g., soil, rocks, and staging equipment) producing spherical shaped

particles of about 0.2-2 mm in diameter; 3) irregularly shaped solid or partially molten
components of ground material; and 4) agglomerates of radioactive particles of types 1 and 2 [4].
The physiochemical properties of particles are shot and device dependent. For instance, spherical
particles are formed at high altitudes whereas ground detonations create vitrified materials with
affixed soil compartments.

The duration and horizontal extent of fallout largely depends on the distance of the detonation
from the ground surface, meteorological conditions, and to a lesser extent the yield of the
weapon [4]. Surface bursts, when exposed to an increased inventory of surface area,  are more
likely to produce local fallout and, therefore, limit the extent of horizontal contamination when
compared to atmospheric bursts. Nonetheless, the effects of nuclear weapons may extend great
distances within a short period of time, and under the right conditions can circumnavigate the
earth in a matter of days [14]. Fallout settling time—or the duration of time it takes for particles
to deposit on the ground—is largely a function  of particle size, as demonstrated in Table 1.

              Table 1. Particle Characteristics According to  Fallout Type [7]
Fallout Type
Local fallout
Intermediate fallout
Global fallout
Large particles
Medium-size particles
Small particles
Radius (um)
> 100
Settling Time
The radioactivity of particles is attributed to fissionable material (i.e., uranium (U) or plutonium
(Pu)) in addition to the generation of fission products. Radionuclides most commonly attributed
to nuclear weapons fallout are [15-17]:

    •   Fission products—cesium-137 (137Cs), iodine-129 (129I), and strontium-90 (90Sr).
    •   Activation products—cobalt-60 (60Co), europium-154 (154Eu), and 152Eu.
    •   Components of nuclear fuel—plutonium-238 (238Pu), 240Pu, and americium-241  (241Am).

2.2. Nuclear Power Plants

NPPs make use of fission to heat water that in return produces steam that actuates steam-driven
turbines. Nuclear fission is achieved by means of U or Pu isotopes [18]. In addition to heat,
fission produces a number of radioactive byproducts. Because of the immense heat generated
during fission, NPP require advanced cooling systems. In the event that these cooling systems
fail, the core containing the fuel and fission products can be compromised. This excessive
heating causes melting of the cores and protective casing (i.e., meltdown). If not remedied,
pressures within the containment structure may exceed safe levels, causing the structure to either
leak radioactive contaminants or explode—dispersing radioactive materials  [18].

Nuclear reactors contain a myriad of radioactive materials. The Chernobyl accident in 1986
released condensed aerosols, fuel particles, and radioactive gases, with the latter contributing
around 50% of the total release. Dobrovolsky and Lyalko assigned Chernobyl particles to four

categories: 1) fuel particles (i.e., fuel particles consisting of U oxide fuel fragments containing a
range of fission and activation products), 2) fuel construction particles (i.e., particles consisting
of a matrix of nuclear fuel), 3) construction particles (i.e., reactor construction materials that
serve as carriers), 4) and hybrid particles or condensation particles (i.e., particles that were
mainly formed from interaction of reactor fuel, construction materials, and fire suppressant
materials) [7, 19]. Two groups of radionuclides were identified within areas contaminated in the
Chernobyl incident:

    •   The volatile group: tellurium-132 (132Te), 134Cs, 137Cs, molybdenum-99 (99Mo),
        ruthenium-103 (103Ru), and 106Ru, with an activity median aerodynamic diameter
        (AMAD) on the order of 1 um.
    •   The refractory group: barium-140 (140Ba), zirconium-95  (95Zr), cerium-141 (141Ce),
        144Ce, 89Sr, and 90Sr, with an AMAD of 4 um [8, 14].

Before Chernobyl, very little 134Cs was present in the environment (i.e., nuclear weapons tests
did not produce a large amount of 134Cs).

Although the effects of Chernobyl fallout had global implications, it was estimated that 2xl018
Bq of fission and activation products were deposited within 30 km of the power plant.
Heightened contamination levels remained within the atmosphere years following initial release
[16]. As with NW detonations, environmental and meteorological conditions and source material
(i.e., fuel) are influential in determining the overall source characteristics. For instance, the
stratospheric residence times of radiocesium were approximately 770 and 150 days for
Chernobyl and Fukushima, respectively [20]. This disparity is likely linked to higher amounts of
refractory elements being dispersed from Chernobyl, and in the case of Fukushima, a reduction
in resuspension because of contaminates being transported  offshore. Similarities were  also
observed. The AMAD of volatile nuclides from Chernobyl was approximately 0.51 um,
compared to 0.43 um for Fukushima [21].

2.3. Radiological Dispersion Devices

RDDs are designed  to disseminate radioactive material in an attempt to inflict casualties, cause
destruction, invoke fear, and ultimately contaminate people and surfaces. The primary effects of
RDDs are twofold. First, although not effective in inflicting mass causalities, RDDs are capable
of exposing bystanders to dangerous levels of radiation. Second,  RDDs are capable of dispersing
and depositing contaminants that will likely require extensive efforts and resources to remediate.

RDDs consist of a dispersal source and radioactive material. The dispersal source will  typically
involve an explosive device; however, this may not always be the case. The yield of the dispersal
source is proportional to the extent of contamination. Radioactive materials are chosen based on
their portability, ease of access, level of activity, and physiochemical forms. The following
radionuclides are of greatest concern: 241Am, californium-252 (252Cf), 137Cs, 60Co, iridium-192
(192Ir), 238Pu, polonium-210 (210Po), radium-226 (226Ra), and 90Sr [3]. It is suspected that the
following factors will be considered by those building RDDs: 1)  availability of a particular

isotope, 2) availability of an isotope in large quantities, 3) ability to shield, 4) high activity level,
and 5) sufficiently long half-life [2]. Literature has suggested that beta/gamma sources between
0.1 and 100 TBq would be best suited for strength and mobility purposes [22]. There have been
300 attempted radioactive substance smuggling incidents over the past decade [2].

RDD composition and geometry can greatly affect the aerosolization properties of the
radioactive material [13]. For explosive detonations, fine particulate matter can undergo a phase
change to either a liquid or a vapor. This will result in lower deposition velocities at the surface.
Particles will remain airborne for longer periods of time and contaminate a larger area, even with
limited explosive force [22]. Upon detonation, two important phenomena take place: 1)
radionuclides agglomerate with inert material inside the fireball, and 2) secondary oxidation of
the radiological material occurs. In the event of agglomeration, small particles of inert material in
the vicinity of the fireball may be entrained into the turbulent eddies within the fireball. These
particles then interact with the radiological material, assisted by varying velocities within the
eddies [13]. As a result of this interaction, radiological particles increase in size, causing rapid
settling. According to Harper, Musolino, and Wente, the opposite was observed when
radiological materials conducive to oxidation (i.e., with a high Bibbs energy of formation) were
used [13]. Particle size decreased significantly resulting in a reduced rate of deposition. RDDs
will likely produce uneven radioactive fragments and particles of varying sizes, resulting in
heterogeneous patterns and areas of contamination [9].  Studies have indicated that particles are
closely associated with nearby materials (i.e., soil matrix). Moreover, these materials influenced
the morphology and chemical composition of particles [23].

RDDs have the potential to contribute a significant amount of exposure; however, the spatial
extent of contamination, by means of explosives, may be limited. The extent and spread of
contamination is dependent upon particle size, height of release, and ambient weather conditions
[24]. Unlike NW detonations and NPP accidents (which can produce either spherical or
fragmented particles, depending on their source and surrounding materials), RDDs will likely
produce particles in a fractured state with moderate particle size distributions. RDDs are less
predictable than NW detonations or NPP accidents in terms of magnitude and extent of
contamination. Where NPP accidents are closely associated with fuel and NW detonations with
the height of detonation, RDDs are a direct product of their construction materials. These
materials will determine the overall physiochemical  characteristics of the contaminants and, most
importantly, the efficiency of aerosolization.


3.1. Deposition

Deposition can be described as the process in which atmospheric material is dispersed through
the air, eventually settling on surfaces [25]. The settling of particles consists of two unique sub-
processes: dry and wet deposition. Dry deposition is the direct interaction of a surface with the
material  excluding the direct incorporation into precipitation [26]. Dry deposition is determined
by a combination of factors such as surface and particle characteristics and meteorological

conditions [8]. During transit, precipitation may lead to a significant depletion of contaminants
[27]. This scavenging process is known as wet deposition. Wet deposition is described as the
transfer of material from the atmosphere to the surface by means of precipitation (i.e., rain or
snow) [25]. Important factors affecting wet deposition are the proximity of the dispersed material
to an area of precipitation, height of the rain-producing layer, amount and rate of precipitation,
and particle size distribution [15, 28-30].

Deposition values may fluctuate considerably due to a number of factors including
meteorological conditions, heterogeneous mixing, and particle size distribution of the passing
plume. This phenomenon was particularly evident following the Chernobyl incident, when areas
near the NPP experienced lower rates of deposition in comparison to outlying areas. Wet
deposition was most prevalent during Chernobyl and was correlated with geographic areas
reporting increased levels of contamination [16, 17]. Surface elevation may also play an
important role in deposition. This is likely due to the presence of orographic precipitation found
at high-altitude locations [31, 32].

3.2. Surface Transport

The migration of contaminants across a given surface is a function of the initial physiochemical
characteristics of the contaminant as well as the composition, orientation, and condition of the
residing surface [25]. These characteristics may vary depending on the environment in which the
contaminant was deposited. The particle's environment (e.g., urban areas) and that
environment's components (e.g., sewers) greatly influence the extent and magnitude of
radioactive contaminant transport. The purpose of this section is to highlight the significance of
urban and rural environments to radioactive contaminant transport as contaminants traverse these
environments via various mechanisms.

       3.2.1.  Urban Environments
Literature originating from Chernobyl and RDD dispersion modeling have shown that dispersion
and deposition in urban areas are largely influenced by the presence of tall buildings that divert
surface airflow patterns around buildings and through city streets [33]. Accordingly, deposition
patterns can vary greatly by city depending on the geometry of the infrastructure [33]. For
instance, following deposition, radioactive material either penetrates into urban surfaces,
migrates by means of fluvial transport, pools in low-lying areas, enters sewer systems, or is
resuspended, restarting the process.

The transport of radionuclides will vary depending on their physiochemical characteristics. For
example, 137Cs, a radioactive isotope of cesium, is highly mobile during periods of precipitation
because of its high water solubility [34]. Surface migration of contaminants due to runoff is
promoted by presence of tall buildings and sewer or drainage systems [25]. Dose rate
information from aerial surveys conducted in the Fukushima evacuation areas have shown a
faster than expected reduction in the overall dose rate [35]. Empirical studies have attributed this
reduction to surface runoff of 137Cs. However, this collective runoff has the potential to create
large areas of pooling containing high concentrations of radionuclides [35]. Decades following

initial deposition from the Chernobyl incident, low-lying urban areas have been found to contain
activity levels as high as 540 Bq/kg when compared to low-lying areas [36]. Very little is known
about the distribution, resuspension, and migration of radionuclides in urban environments [37].
Despite this, their ability to expand contamination in an area has been well documented in
literature [37].

       3.2.2. Buildings
Buildings account for a majority of the surface area within urban environments. Depending on a
number of factors, buildings have the potential to collect or promote the migration of
radionuclides. As with the radionuclide itself, this is highly dependent on the physiochemical
properties of the building construction materials. Studies assessing the distribution of
contaminants from Chernobyl in nearby urban areas have suggested that building orientation in
regard to release point and building height may influence the distribution of radionuclides on
building surfaces, with the upper floors containing higher levels of contamination [37]. This is
likely a result of inadequate shielding of nearby buildings of increased heights coupled with the
effects of urban airflow. However, to what extent these characteristics influence the deposition
and migration of radionuclides is still unknown.

       3.2.3. Streets and Sidewalks
Levels of contamination of the streets of Pripyat, an urban area approximately 100 km from
Chernobyl, were an order of magnitude or more lower than in buildings [33,38]. Contamination
levels of streets are known to rapidly decrease over time because of weathering and mechanical
agitation; however, less traveled streets tend to retain contaminants  for longer periods of time
[34, 38]. One study in particular reported contamination remaining on roadways for up to 7 years
[34]. A majority of the contamination resides on a thin layer of street dust and is removed as a
result of resuspension or migrates to neighboring surfaces by way of runoff [34].

Due to the type of materials commonly used in their construction and their proximity to
buildings and streets, sidewalks present activity levels similar to streets following a radiological
incident [39]. Cement, a material often used in the  construction of sidewalks, is conducive to the
entrapment of radionuclides, such as 137Cs, due to the presence of silica and the micro-porosity
of the cement matrices [39]. Because of this, sidewalks may retain contaminants for longer
periods of time than other urban surfaces.

       3.2.4. Sewers
Following initial deposition, wash-off mechanisms such  as runoff from precipitation or wash-
water from decontamination activities are capable of transporting radionuclides throughout urban
areas. This phenomenon is augmented by the presence of drainage systems and the increased
distribution of impermeable surfaces [25]. Contaminated effluents subjected to runoff by means
of wash-off will either 1) migrate through urban drainage systems, risking redistribution in water
courses elsewhere, or 2) enter sewer treatment systems where they may be trapped in sewer
sediment or sludge, treated and disposed of, or recirculated for reuse [25]. It is assumed that the
transport of radionuclides is largely dependent on the species, physiochemical characteristics,
construction material, and status of sewerage systems. Species may be the most influential factor.
In two separate  studies, a mass balance of effluence entering the sewage treatment plant showed

10% of 137Cs and 20% of 90Sr being trapped in sewage sludge, with the remaining fraction being
discharged [40, 41]. Urban drainage systems and treatment systems are likely the most efficient
mechanisms at transferring contaminants to clean areas or exposing inhabitants to moderate
levels of radiation. Nonetheless, very little information exists on the fate of radionuclides in
sewer systems.

       3.2.5.  Parks
Gardens and parks tend to act as sinks for radiological contaminants in urban environments and
are widely reported as contributors to the overall dose [42, 43]. This is likely due to the
abundance of permeable surfaces (e.g., soil) within these areas. This phenomenon may be
augmented by runoff from surrounding impermeable urban surfaces. Surface transport
mechanisms in gardens and parks are similar to those found in rural environments.

3.3. Rural Environments

Permeable surfaces are more prominent in rural areas than urban ones, due to the abundance of
soil matrices. Accordingly, either radionuclides are captured or migration is greatly reduced.
Nonetheless,  rural environments will  likely retain radionuclides for extended periods of time, act
as sources of radiation, or contaminate foodstuffs.

       3.3.1.  Forests
Forests near Chernobyl and Fukushima greatly reduced the overall mobility of radionuclides [44,
45]. Forests with litter layers present tend to not only slow migration, but also reduce levels of
resuspension [44].  Over time, forests  tend to homogenously distribute contaminants throughout
the ecosystem. Very little radioactive 137Cs is lost from the system by runoff, with more than
90% of the total  deposition residing in the upper organic horizons. The underlying litter layer
constitutes as a major regulator of forest systems [46]. Given their use for resources, natural
habitats, and recreation, and their proximity to populated areas, forests should be a major
deciding factor when considering decontamination.

Trees are a predominant plant of forest ecosystems and are extremely efficient at collecting
pollutants [46]. Dry deposition on trees may be 10 times higher than in adjacent grasslands [9].
Scavenging efficiency and subsequent migration is a function of species, season, and
environmental conditions [27]. Scavenging efficiency greatly varies by species. For instance,
conifers have a greater interception capacity than broad-leaved deciduous trees. Upon deposition
trees may translocate radionuclides from the underlying soil  matrix or the atmosphere.
Contaminants are either absorbed by leaves and needles or migrated by means of root uptake or
translocation [47].  Younger trees may lack the more extensive canopy afforded to older trees
[48]. Therefore, the translocation of contaminants for small trees is less likely. The absorption of
contamination by root uptake has been described as limited due to the retention qualities of soil.
Emerging tissue contains  very little radioactivity; this suggests that it  is not easily absorbed and
transferred to other tissue, with a large portion residing in the needles, twigs, and leaves versus
the core wood [49]. Over time, contamination will continue to accumulate in the core wood [50].

Water availability is described as a major factor in determining retention rates (i.e., more water,
higher retention) [49].

Contamination of trees was found to be roughly three times higher than that of grassed surfaces.
One study in particular reported forest canopies intercepting as much as 60%-80% of
radionuclide deposits [27, 44]. This is likely due to the abundant surface area of leaves/needles
[27]. Contamination eventually migrates to the litter layer by means of runoff during periods of
rain or leaf shedding [44]. Radionuclide redistribution in a forest canopy may be enhanced by the
following mechanisms: 1) removal via wash-off or wind erosion; 2) foliar penetration or root
uptake; and 3) leaf/needle removal [48]. Wash-off was found to account for approximately 80%
of migration [51]. When tree growth is active (i.e., spring, summer), the removal of contaminants
found in the forest canopy is accelerated. This acceleration is likely due to shedding of epidermal
layers of leaves and bark. This phenomenon typically occurs within the first couple of warm
months [44].

Literature once presented the hewing of heavily contaminated areas of land as a productive
countermeasure [44]. The reduction of activity by pruning and removal was later proven to be
inefficient [52]. It is now assumed that leaves and branches may act as shields of radiation,
especially for urban areas adjacent to dense forests [52]. However, the removal of topsoil,
immediately following deposition, was found as a  productive countermeasure in reducing the air
dose rate [52]. There may be areas in which public perception may require forests to be culled, as
in Chernobyl's "Red Forest" [44]. If hewing of large portions of forest areas is desired, it is
highly recommended that these activities take place during winter months preferably when snow
cover is present, in order to limit the disturbance of the litter layer [44].

       3.3.2.  Soil
As a result  of NW testing and nuclear NPP accidents (e.g., Chernobyl), the migration of
radionuclides in soils is well documented throughout literature.  Studies have consistently
described soil as a major pathway for radionuclides entering the food chain and water supply.
Furthermore, the contamination of soil has the potential  to retain radiological materials for
extended periods of time and to transfer contaminants to otherwise clean areas. Radionuclides'
vertical migration is gradual. It is reported that some radionuclides take years to migrate just a
few centimeters [53]. The rate of migration varies by radionuclide. The most commonly studied
radionuclides according to rate of migration in soil are as follows: 131I > 239Pu, 240Pu > 90Sr >
137Cs [5, 15, 34, 42, 53-58]. Table 2 below shows the mean depth profiles for 240Pu and 239Pu,
90Sr, and 137Cs from various radiological sources.

        Table 2. Mean Depth Profiles for 240Pu and 239Pu, 90Sr, and 137Cs [53, 59-61]
240 + 239pu
Nuclear weapon
Mean Depth
0-2 cm
5-10 cm
0-5 cm
0-5 cm
0-5 cm
Years Since Deposition
>50 years
14 years
<1 year
12 years
<1 year
A number of variables influence to the mobility of radionuclides in soil. The most commonly
attributed are soil solution, physiochemical properties of soil, microorganisms, the radionuclide
involved, and moisture level [42, 47, 56, 58, 59, 62]. Accordingly, lower pH-values, clay
content, and cation exchange capacities result in higher mobility rates [47, 56, 59, 62]. Migration
is most apparent when contrasting ecosystems and climates. Radionuclides such as 137Cs have a
stronger affinity for semi-natural ecosystems than for agricultural ecosystems, with forest
ecosystems having a prolonged rate of attenuation [12]. Radionuclide efflux is minimal during
winter months and at its maximum during summer months. This is likely due to the increased
presence of biomass in soil during winter [63].

       3.3.3.  Aquatic Environments
Effluent from the source to surface waters is an important pathway for further contamination and
exposure (e.g., drinking water, irrigation systems) [64]. Migration is typically facilitated by
wash-down and surface runoff. Groundwater contamination resulting from rainwater infiltration
from soil is not expected. Rather, considering the slow migration rate of most radionuclides, the
odds of exposure from other environments are higher [65].

Upon entering aquatic environments, radionuclides accumulate in bottom sediments and organic
matter [71]. Factors influencing level of contamination include subsurface contamination levels,
time of inundation, and volumetric flow rate [66]. Contamination levels decrease in a relatively
short period of time by means of flushing, burying, and radioactive decay [12]. Attenuation is
most noticeable in oceans, rivers, and lakes (with tributaries); closed areas of water tend to
decrease at a slower rate.  This is likely from the lack of discharge resulting in minimal
sedimentation (i.e., burying).

Contaminants may be remobilized due to precipitation and flooding events [67, 68].
Concentrations are closely correlated with increasing volumetric flow rate. Flow rate is a
function of precipitation and flooding. As water flow and water levels increase, so does the
likelihood of remobilization [67]. Many studies have demonstrated that this phenomenon,
particularly for 90Sr and 137Cs, is closely correlated with flooding events [69]. River banks and
flood plains near large rivers have been of particular concern with regard to the Fukushima
incident. These areas appear to collect high concentrations of 137Cs following significant

precipitation events and continue to act as a transport mechanism years following initial
deposition [35].


Resuspension is defined as the re-entrainment of atmospherically deposited material into the
atmosphere [70, 71]. Resuspension is the primary source of lateral spreading of radioactive
material outside the initial contaminated area [72]. This lateral spreading serves as a persistent
source of both dose and contamination. Inhalation of resuspended particles, within the first few
weeks of initial deposition, may have the same impact as directly inhaling the initial
contaminating cloud [73]. Literature has consistently defined resuspension as a potential concern
following a radiological release. During Chernobyl, recently cleaned areas were often re-
contaminated,  regaining initial levels of contamination [74]. Following the end of weapons
testing, the primary mechanism leading to increased atmospheric concentrations was the
resuspension of previously deposited material; this is still considered the most significant
contributor for both air and ground activity [11,  17].

Historically, the effects of resuspension were not considered a major threat [75]. During nuclear
weapons testing in the 1950s and 1960s, the dose of field observers were drastically
underestimated as a result of resuspension [73]. What little data resulted from these assessments
originated from tests conducted in arid regions where there was a strong dependence on
resuspension by wind [76]. Furthermore, historical resuspension measurements were often
delayed until months or years after the initial event. Until the advent of Chernobyl, there had
been a limited number of studies investigating resuspension  [77].

Literature has defined a number of factors that may influence resuspension. The most prominent
are 1) surface properties,  2) meteorological conditions, 3) size distribution of contaminant
particles, 4) chemical properties of contaminant, and 5) time since initial deposition [62,  70, 76,
78-82]. Particles are more likely to resuspend when attached to a (larger) host particle [74, 83].
The act of resuspension may be initiated by a number of mechanical influences [77].
Resuspension can be initiated by human or natural mechanisms [70, 82]. Figure 1 below shows a
list of common resuspension mechanisms according to natural and human activities.

     Figure 1. Resuspension Mechanisms According to Natural and Human Activities

4.1. Natural Resuspension Causing Mechanisms

       4.1.1.   Wind
The influence of wind on resuspension in not certain, and literature is conflicting [84]. A
majority of studies have indicated that resuspension will increase with wind speed and particle
size (i.e., area), which in return promotes aerodynamic lift [72, 75, 82, 85, 86]. However, a few
studies found correlations with wind direction instead [76, 78]. For wind-driven resuspension,
particles of 1,000-2,000 um roll or slide across a surface depending on wind direction (i.e.,
surface creep). Particles smaller than 50-1,000 um in diameter may be lifted by wind and
subsequently return to earth as a result of saltation [86, 87]. Once the particle strikes the ground,
it may resuspend smaller particles  [72,  87]. Studies have shown that most radioactive material
transported by wind is a result of saltation [86]. Under similar conditions, particles less than 50
um may be suspended for extended periods of time. The resuspension of small particles less than
0.1 um is unlikely. Nonetheless, over time, contamination levels are rapidly diluted in the
presence of heightened wind speeds [75, 85].

It has been historically assumed that trees may act as  sources of resuspended material secondary
to wind transfer [79]. Early studies out of Chernobyl indicated that,  once contaminated, trees
could serve as a spreading mechanism. It was hypothesized that vegetation induces resuspension
due to the friction of leaves and disturbances in turbulence [72]. Recent literature indicated that
contaminants deposited on tree canopies (specifically of coniferous trees) near Chernobyl
remained for an extended period of time. Furthermore, some studies showed higher levels of
contamination at the edge of forests, indicating that trees near the forest's edge were, in effect,
capturing a majority of contaminants before  they entered the forest [44, 48]. The rate of
resuspension of small particles from surfaces of trees due to wind has been described as minimal

and decreased over time [79]. In fact, the presence of vegetation may minimize the effects of
resuspension and reduce atmospheric concentrations of radionuclides over the long term [79, 88].

       4.1.2. Rain
Literature suggests rain is capable of causing both mechanical and surface agitation, which can
lead to resuspension. This phenomenon is temporary: as rain saturates the surface, the rate of
resuspension is greatly reduced. This mechanism has been described as a largely ineffective
resuspension mechanism [80, 89].

       4.1.3. Fire
Following a radiological incident, access to forested areas will likely be restricted, significantly
reducing human influence on wildfires. Rather, natural mechanisms such as lightning are more
likely to cause wildfires. Therefore, for the purposes of this report, wildfires are deemed as a
natural mechanism of resuspension.

Biomass, particularly in forested areas, is extremely efficient at collecting radionuclides by
means of aerosol capture or absorption. Biomass may retain high levels of radiation over many
years  or decades. The presence of combustible vegetation increases the risk of fire in any
forested area; literature has shown that radionuclides may resuspend during forest fires  and are a
significant inhalation hazard [90, 91]. This is particularly true in the presence of alpha emitters,
where exposures can quickly exceed safe levels [90]. During burning, atmospheric
concentrations may increase by several orders of magnitude [92].

Atmospheric concentrations are a function of a fire's phase. Concentrations can be  several
hundred times higher during active burning,  up to 10 times higher when a fire is smoldering, and
up to  several times higher post-fire [90]. Accordingly, the rate of resuspension is a function of
fire temperature and local weather conditions [46]. The burning process may modify the
aerodynamic properties of particles, thus increasing their risk of resuspending [90]. A study
conducted by Horrill, Kennedy, Paterson, and Mcgowan estimated that between 10% and 40% of
radiocesium was released from forest fires near Chernobyl [46].  Moreover, a recent study by
Mousseau, Milinevsky, Kenney-Hunt,  and M011er suggested that, in highly contaminated forest
areas, litter loss may be depressed by reductions in soil invertebrates, thereby increasing the risk
of forest fires [93]. Overall, the range of particles resuspended due to forest fires is  limited.
Estimates range between several kilometers  and up to 20 kilometers [90, 92].  Although
concentrations of radionuclides are high in close proximity,  fire events will only impact a very
small  area downwind [92].

4.2. Human Resuspension-Causing Mechanisms

Anthropogenic resuspension has been associated with a wide range of activities [76, 78, 82].
These can reach greater depths, and thus resuspend a larger number of particles [87].
Historically, the effects of resuspension in urban environments are poorly understood, though
Chernobyl and Fukushima brought marginal advancements in understanding [75].

       4.2.1.  Vehicles
The resuspension of material from roadways could lead to increased inhalation dose and serves
as an important factor in the (re)contamination of localized areas [94]. Literature addressing the
resuspension of radionuclides in urban environments as a result of vehicle activity is severely
lacking. What limited work has been conducted clearly indicated that material could be
resuspended by vehicle traffic. This was demonstrated in a widely cited study by Garland, who
correlated high concentrations of resuspended pollutants during periods of increased vehicle
activity [76].

Vehicles primarily cause resuspension through particle abrasion by tire sheer and induced
turbulence [70, 72]. The resuspension of particles from vehicle traffic is highly dependent upon
particle size, vehicle size, road surface type, time since initial deposition, and vehicle speed, with
the latter being the biggest contributor [74, 76, 80, 94-96]. A study by Nicholson, Branson,
Giess, and Cannell showed that the rate of resuspension increases with particle size [94].
Resuspension from vehicle-induced turbulence requires that the threshold value exceed the
surface stresses on the particle [95]. This threshold value is determined by the surface area of the
passing vehicle (i.e., the larger the vehicle, the greater the displacement of air); therefore, trucks
will likely resuspend more material than cars. Vehicle speed  is also an import factor. As vehicle
speed increases, so does the  rate of resuspension (i.e., proportional to turbulence generated by the
vehicle). The difference in vehicle surface area is less evident at speeds above 20 mph [96].

Roadway surface type is a major factor in  determining the surface particle matrix. For instance,
particles may behave differently on dirt roads due to the increased presence of dust particles that
may act as carriers [95]. Furthermore, literature has indicated a reduction in resuspension when
vehicles were driven over vegetative surfaces rather than paved surfaces [74, 96]. Particles
residing on roadways have a fairly short residence time in that a majority of the material is
displaced within the first few vehicle passes. This is likely due to a reduction in particle
inventory resulting from resuspension or fixation of particles to the road surface. Literature is
conflicting in the amount of time for which material is readily resuspended. Estimates range
from five days to a year [76, 80, 95, 96]. Generally, the rate in which material is resuspended is
reduced by an order of magnitude within the first week, and several orders of magnitude within
30 days [80].

There are a number of outlying risks associated with vehicle-induced resuspension. In some
instances, larger particles may be resuspended without the tire itself being introduced directly to
the material [70, 74]. Furthermore, once mechanically agitated, particles may separate into
smaller sub-particles, possibly increasing their atmospheric residence time [81]. Although most
prominent in dry conditions, tire spray was reportedly a significant contributor during wet
surface conditions and is capable of transferring particles to adjacent surfaces or passing vehicles
[74, 94]. Vehicles are likely  the primary source  of resuspension in urban areas. Vehicle activity
should be limited following initial deposition at the risk of remobilizing contaminants and thus
increasing the extent of contamination.

       4.2.2.  Pedestrians
Very little literature on pedestrian-driven resuspension exists. A study by Linsley showed that
bicycle and foot traffic caused more resuspension than wind, but only about a quarter as much as
vehicle traffic [96-98]. Overall, pedestrian-generated resuspension is not considered a significant
contributor of radiological particle resuspension.

       4.2.3.  Agricultural Activities
Literature is conflicting as to the degree of resuspension generated by agricultural activities. A
number of studies have assessed particle resuspension as a result of Chernobyl and Fukushima.
They collectively found that agricultural practices such as tilling have the potential to increase
atmospheric dust levels and concentrations of radionuclides [82, 84, 99]. However, the
resuspended particles were very limited in lateral spreading. This is likely because of
agglomeration, in which smaller particles affix to larger particles originating from the soil
matrix. Only those individuals operating farm machinery or within close proximity of
agricultural areas would be at risk. Resuspension because of agricultural activities is  considered a
moderate spreading mechanism, capable of redistributing material up to a few hundred meters

       4.2.4.  Decontamination Activities
Decontamination activities have shown to significantly increase radionuclide concentrations,
particularly downwind [88].  The rate at which particles are resuspended is dependent on a
number of variables including decontamination technology being used and meteorological
conditions. More research is needed to determine the impact of resuspension as a result of
decontamination activities; however, preliminary Fukushima studies suggest that nearby outdoor
decontamination activities correlate with increased concentration levels of 137Cs in indoor areas

4.3. Factors Influencing Resuspension

       4.3.1.  Particle Size
It is suspected that resuspension rate is a function of particle size due to an increase in
aerodynamic lift force that increases with cross-sectional area [80]. Nonetheless, small particles
are less subject to gravitational forces and may be carried over long distances. For instance,
particles that originated from NW testing can typically be divided into three different categories:
1) large particles that are deposited onto the surface within a matter of hours, 2) small particles
that remain in the atmosphere for days, 3) minute particles (i.e., the remaining fraction) that can
remain in the atmosphere for months to years [28]. Particles 10 um and smaller are subject to
atmospheric transport, whereas particles 20 um or larger are deposited near the source [101].

       4.3.2.  Range
The range of resuspended  material is fairly limited [76, 78, 96]. Literature has reported a few
instances where resuspended material traveled long distances (i.e., 50-60 km); however, very
little deviation of contaminated areas has been reported [75]. For instance, while characterizing
contaminated areas in Goiania, Brazil, following an accidental release of 137Cs, Pires Rio and


Paretzke noted that the remaining activity was still confined to within 200 meters of the initial
deposition site [102]. In this location, resuspension did not contribute to the overall exposure.
Significant contamination of agricultural products, equipment, and structures was reported
however. Furthermore, a  study by Garland indicated that 60% of particles between 10 and 30 um
were removed from grass and redistributed within 4 meters by means of wind-driven
resuspension [80, 89]. Regardless of particle size, literature disagrees in terms of range.
Estimates vary from a few meters to up to 60 kilometers [76, 80, 89]. This disparity is likely a
result of variations in climates across research locations.

       4.3.3.  Timeline
Resuspension will be particularly high during the early phases of a radiological incident [22, 72,
81]. These levels decline  linearly with time. As previously deposited small particles are removed
by weathering and resuspension, only those particles that are firmly bound to the surface remain
[103]. Areas contaminated some 10-20 years ago may still serve as a primary source  for
resuspension; however, the majority of available literature is in agreement that resuspension may
be possible at least three years following initial deposition [70, 76-78, 94, 102].

       4.3.4.  Limiting Factors
Ignoring the physicochemical properties of the particle and the surface-substrate, there are a
number of factors that influence the availability  of particles, at the surface, for resuspension.
These variables vary by location. The most prominent are as follows:

   •   Weathering: As a result of weathering, particles are either captured by sorption or are
       washed away [98, 104]. The rate of resuspension is greatly reduced after the first
       precipitation event [72].
   •   Displacement: Both human and natural disturbances may transport contaminants to
       sheltered locations that are less prone to resuspension [98].
   •   Migration: The vertical migration of radionuclides, albeit slow, will ultimately reduce the
       amount of particles available for resuspension in the distant future and may act as  a
       shielding mechanism as the contaminant penetrates to greater depths  [35, 59].
   •   Proximity: Larger particles have a higher likelihood  of depositing near their source and
       are more susceptible to resuspension; smaller particles, located near the periphery, are
       less susceptible to resuspension [76].
   •   Time: The rate of resuspension declines linearly with time [98].
       4.3.5.  Climates
Climate is a major factor  in resuspension. In arid climates, resuspension is prominent by virtue of
decreased precipitation, low humidity, and minimal vegetation [75,  105]. In humid climates with
pronounced vegetation and precipitation, resuspension may be limited [80].

       4.3.6.  Seasons
There is a noticeable uptick in atmospheric concentrations of radionuclides during the spring and
summer months  [70]. Peaks in concentration in autumn and winter can be attributed to a limited
vegetation matrix [75]. Increased wind  speed during spring and autumn may also increase the
rate of resuspension [75].



Computational methods for estimating the effects of radiological material on people and the
environment originated from NWs testing in the 1950s and 1960s [106]. These models, although
simplistic in nature, were adequate for predicting the homogenous spreading of radionuclides
produced by NW detonations and were well suited for the terrestrial geometries presented by arid
environments. The Chernobyl  era ushered in a need for more complex computational methods
for modeling the spreading of  radionuclides outside arid environments [106]. Predicting
exposure pathways became increasingly complex when considering urban geometries: unlike
arid environments, urban areas aid in the heterogeneous spreading of contaminants, which
required complex  computational codes. Nevertheless, as Chernobyl beckoned international
attention, significant intellectual and financial resources followed [9, 106]. More capable
calculation methods soon became available (e.g., Monte Carlo simulations).

Today, decision-makers can exercise a variety of tools for modeling the dispersion, transport,
and overall fate of radionuclides for the purpose of predicting exposure, assessment, and
remedial activities [9]. In the absence of monitoring or sampling measurements, these tools
enable decision-makers to assess current or future conditions or implications according to a set of
established parameters or environmental conditions. With the occurrence of the Fukushima
accident, modeling capabilities are again challenged with the need for increased spatial scale.
Population densities and overall  urbanization are higher than in the Chernobyl area, requiring
more precise modeling capabilities that can account for the effects urban environments and their
surrounding rural  areas might  impose on particle transport.  Therefore,  as with the post-
Chernobyl period, an assessment of current technology is necessary in order to determine the
current capability  to model small- and large-scale radiological incidents in urban areas. Although
a high-level assessment of radiological models is beyond the scope of this report, an inventory of
publically available radiological  models was conducted in order to assess their current state so
that recommendations pertaining to future research can be made. A literature review was
conducted to identify the most widely cited radiological models. Sections 5.1-5.8 provide an
overview of those models.

5.1. ARGOS

Summary: ARGOS is an information system (IS) for enhancing crisis management for incidents
involving chemical, biological, radiological,  and nuclear (CBRN) releases. ARGOS was
designed to address a range of CBRN incidents. ARGOS is a prognostic tool as well as a
database system for the collection and presentation of data for use in emergencies in an easily
understandable form.  ARGOS facilitates decision support, improving of situation awareness, and
information sharing among emergency response organizations. As a simulation instrument,
ARGOS is a valuable training  tool for use by emergency response organizations.

Affiliated organization: Prolog Development Center A/S,  Ris0 National Laboratory, Danish
Emergency Management Agency

Model type: Decision support system

Model algorithm:

   •   Source model: Initially developed by Ris0, the Source Calculation Model predicts gas
       releases from industrial storage or transport containers. The Source Calculation Model
       also calculates the initial dispersion until a stage where other dispersion models of the
       ARGOS system are able to take over.
   •   Atmospheric dispersion: ARGOS is divided into a small- and mid-scale dispersion model
       (called LSMC/REVIPUFF), a model for dispersion in urban environments, a model for
       dispersion of heavy gases, and a system for coupling to long-range models: DERMA
       (Danish), MLDP (Canada), SNAP (Norway), and MATCH (Sweden).
   •   CBRN scenarios: ARGOS incorporates an urban dispersion model capable of accounting
       for urban and street canyon geometries.
   •   Nuclear scenarios:  ARGOS includes a module for simulating the transfer of radioactive
       material in food chains, and for assessing doses via all relevant pathways (i.e., internal
       exposure via inhalation and ingestion,  external exposure from the plume and from
       deposited radioactive material) to the population. The food dose model estimates doses to
       the public (individual as well as collective doses).
   •   Countermeasure modules:
          o  The Agricultural Countermeasure Program (AGRICP), which extends the food
             dose model to cover agricultural countermeasures, instead of just calculating
          o  The STRATEGY food-chain Countermeasure model, developed under the
             European Commission's Fifth Frame-work Program.

5.2. ERMIN

Summary: The European  Model for Inhabited Areas (ERMIN) was developed under
EURANOS, an integrated  project of the European Commission Sixth Framework Program.
ERMIN is both a model and a software tool. As a model, it simulates the behavior of
radionuclides in the inhabited environment and calculates the exposure of the population as well
as other relevant endpoints. ERMIN brings together a number of models and datasets and
embeds an actual transfer model that also takes into account the weathering of material on
building surfaces and movement of radionuclides around inhabited environments. As a tool it
allows a user to explore different recovery options following the contamination of an urban
environment with radioactive material. ERMIN was designed to be implemented within both the
RODOS and ARGOS Nuclear Emergency Decision Support System, and also as a standalone

Model type: Decision support system

Affiliated organization: Health Protection Agency, Centre for Radiation, Chemical and
Environmental Hazards, UK (Now Public Health England); Ris0 National Laboratory for


Sustainable Energy, Technical University of Denmark, Denmark; Helmholtz-Zentrum
Muenchen, Germany-Forschungszentrum Karlsruhe, IKET, Germany; Danish Emergency
Management Agency, Denmark; and Prolog Development Center, Denmark-Bundesamt fur
Strahlenschutz, Germany.

Model algorithm:

   •   Ratios to distribute deposition on the reference surface onto all urban surfaces.
   •   Databases for estimating indoor and outdoor dose rates.
   •   Convection-diffusion equation for estimating migration.
   •   Empirical functions for estimating long-term retention and resuspension.

5.3. RODOS

Summary: The Real-time On-line DecisiOn Support (RODOS) tool was designed as a
comprehensive system, incorporating models and databases, for assessing, presenting, and
evaluating accident consequences over all distances, taking into account the mitigating effects of
countermeasures. The flexible coding used to develop RODOS enables it to cope with
differences in site and source term characteristics, in the availability and quality of monitoring
data, in national regulations and emergency plans, etc. RODOS provides decision support on
four levels:

   1.  Acquisition and checking of radiological data and their presentation, directly or with
       minimal analysis, to decision-makers, along with geographic and demographic
   2.  Analysis and prediction of the current and future radiological situation (i.e., the
       distribution over space and time in the absence of countermeasures) based upon
       information on the  source term, monitoring data, meteorological data, and models.
   3.  Simulation of potential countermeasures (e.g., sheltering, evacuation, issue of iodine
       tablets, relocation, decontamination, food bans), in particular the determination of their
       feasibility and quantification of their benefits and disadvantages.
   4.  Evaluation and ranking of alternative countermeasure strategies by balancing their
       respective benefits  and disadvantages (e.g., costs, averted dose, stress reduction, social,
       and political acceptability) while taking account societal preferences as perceived by

Affiliated organization: European Commission

Model type: Decision support system

Model algorithm: RODOS consists of three primary modules, each with numerous sub-modules
performing a different function (i.e., modular). The primary modules include:

   •   DIsPersion over COmplex Terrain (DIPCOT), an atmospheric  dispersion  model using a
       Lagrangian puff/particle methodology based on a Langevin equation.
   •   A hydrological module.


   •   A terrestrial food chain module and a terrestrial dose module.
   •   An aquatic food chain module and an aquatic dose module.
   •   A forest food chain and dose module.
   •   A tritium food chain and dose module.
   •   A dose combination module.
   •   A countermeasure  subsystem (i.e., early and late countermeasure models).

5.4. CONDO

Summary: The CONsequences of Decontamination Options (CONDO) tool was developed for
estimating the consequences of decontamination options, including dose, required resources,
timescales, and waste management. CONDO incorporates the EXPURT external dose model for
inhabited areas.

Affiliated organization: UK Environment Agency (EA) and Food Standards Agency (FSA) for
version 2.1; EA for version 3; research contractor: National Radiological Protection Board

Model type: Decision  support system

Model algorithm: CONDO consists of a factor-based deposition, radionuclide distribution, and
decontamination model. The dose model (i.e., EXPURT) calculates external whole-body dose
and effective dose from resuspension. External dose to the public is calculated by summing
calculated exposures from different surfaces for varying exposure times and shielding factors.
Doses in various time periods are supplied by a database; external doses are calculated at run
time, and are subsequently integrated from time since deposition to time of interest, accounting
for decontamination.

5.5. MOIRA

Summary: The MOdel based computerized system for management support to Identify optimal
remedial strategies for  Restoring radionuclide contaminated Aquatic ecosystems (MOIRA)
software system is a decision support tool that allows users to evaluate optimal intervention
strategies for radiological contamination of aquatic ecosystems. The software uses various
mathematical models to assess the transport and radiation dose of 137Cs and 90Sr and the effects
of various countermeasures. The software uses Multi-Attribute value Analysis (MAA), which
allows evaluation of various countermeasure alternatives by accounting for impacts on the
economy, society, and  the environment.

Affiliated organization: European Commission

Model type: Decision  support system

Model algorithm: The dose models are based on dose rate factors. The concentration models are
composed of differential equations.


Summary: RESidual RADioactive (RESRAD) is a computer code developed at Argonne
National Laboratory to calculate site-specific residual radioactive material guidelines in addition
to radiation dose and excess lifetime cancer risk to an onsite resident (i.e., a maximally exposed
individual or a member of a critical population group). Major pathways include:

   •   External radiation exposure.
   •   Internal radiation dose from inhalation (dust and radon).
   •   Internal radiation dose from ingestion:
          o   Drinking water (surface and/or groundwater)
          o   Produce, meat, and milk
          o   Fish
          o   Soil

Affiliated organization: Argonne National Laboratory

Model type: Risk assessment

Model algorithm: Mass balance is maintained between the contaminated source and each
transport pathway. RESRAD tracks losses from radioactive decay, leaching (sorption-desorption
ion exchange), erosion, resuspension, and volatilization.

5.7. MOGRA

Summary: Migration Of Ground Additions (MOGRA)  is a code to predict the migration of toxic
substances in terrestrial environments, including radionuclides. The computational code consists
of a dynamic compartment model, a graphical user interface for model formation, computation
parameter settings, and results displays. MOGRA also includes various databases for
radionuclide decay data;  solid and liquid distribution coefficients; soil and plant transfer factors;
feed, beef, and milk transfer coefficients; concentration factors;  and age-dependent dose
conversion factors for many radionuclides. Additional code MOGRA-MAP can import GIS files
and calculate target land  areas.

Affiliated organization: Japan Atomic Energy Research Institute (JAERI)

Model type: Dynamic compartment model

Model algorithm: Simultaneous solution using six-step fifth-order Runge-Kutta method, or


Summary: Program for the Assessment of RAdiological consequences in a Town and of
Intervention after a radioactive contamination (PARATI) is a dynamic model developed to assess


the long-term consequences of accidental radiological contamination of urban environments. The
model is based on empirical data from the Chernobyl and Goiania accidents.

Affiliated organization: Institute de Radioprote9ao e Dosimetria, Brazil

Model type: Dynamic compartment model

Model algorithm: Surface activities and contributions to exposure fields


6.1. Source Characterization

Regardless of the source and explosive yield, high-magnitude radiological events such as NW
detonations or NPP accidents will  likely generate a large amount of radioactive materials. The
physiochemical properties of these materials will depend on a number of variables present at the
time of release. Deposition, resuspension, and concentration values resulting from NPP accidents
and NW detonations are within one order of magnitude. Furthermore, NPP accidents and NW
detonations are very similar in scope, in that the area of contamination following initial
deposition is substantial in size and remains largely unaltered over time. Any sizeable
disturbances of deposited material will quickly be diluted or will pale in comparison to the extent
of initial deposition. Furthermore,  the nature and magnitude of contamination for NPP accidents
and NW detonations are comparable. To the contrary, even under the most pristine conditions,
RDDs are limited in terms of magnitude. Like NPP accidents and NW detonations, RDDs are
susceptible to particle transport; however, given the minimal distribution of contamination
resulting from RDDs, the effects of transport mechanisms are more apparent. For instance, the
Bravo shot of Operation Castle,  a  15-megaton nuclear weapon test conducted in 1954,
contaminated over 11,000  square kilometers [4]. In contrast,  the majority of highly radioactive
material released by an RDD will likely be limited in lateral spreading (i.e., a few city blocks)
[107]. It is for this purpose that this paper highlights the extent of contamination as the most
significant disparity, in terms of particle transport, between large-scale (i.e., NW and NPP)
incidents and small-scale radiological sources (i.e., RDDs).

6.2. Surface Migration

Surface migration plays an important role not only in the dose received by local inhabitants, but
in the overall fate of radionuclides. Once deposited, radioactive particles may migrate by means
of weathering. The rate of migration is largely dependent on the physiochemical properties of the
contaminant and on the surface in  which it resides. The physiochemical properties will vary
depending on a number of factors  contributed to the source of contamination. Furthermore, the
residing surface, in particular its physiochemical properties, orientation, and condition, are
important components of surface migration. These characteristics may vary greatly across urban
and rural areas. Urban areas, for instance, boast an extensive surface area that is oriented
vertically or horizontally, a majority of which is less permeable than rural surfaces. During


precipitation events, these less-permeable surfaces give rise to fluvial transport. Particles then
collect in low-lying areas or are transported into sewer or drainage systems, where they are either
captured or dispensed into natural aquifers [25]. This phenomenon has been noted in Fukushima
studies where the mobilization of 137Cs in urban areas is significant in comparison to undisturbed
flat fields [35]. In contrast, rural areas consist largely of permeable surfaces that tend to act as
sinks for radioactive contaminants. This phenomenon is particularly noticeable in soil where
migration rates of radionuclides are reduced and may remain for many decades.  This migration
process by runoff is best indicated by a runoff coefficient—a dimensionless factor for estimating
the amount of runoff resulting from rainfall. This factor varies depending on the type,
orientation, and level of saturation of a particular surface [108]. Runoff coefficients vary greatly
by environment. Water movement on urban surface is easier than rural surface. In terms of
mobility, this phenomenon might increase the area of distribution of radionuclides—especially
water soluble ones. For instance, the runoff coefficient for an urban area is between 0.30 and
0.50, whereas values for forest values are much less (i.e., 0.05-0.20) [108].

Literature has shown that rural or natural environments are somewhat efficient in managing
pollutants and are of less concern, as they are not likely to impact a large number of people. This
is where the distinction between urban and rural environments in terms of surface migration is
most evident. While rural areas are efficient in capturing contaminants, urban environments are
designed to reroute runoff to outlying areas away from populated areas. In regard to a
radiological incident, this design is rather adverse and is prone for spreading contaminants to
otherwise clean areas. Therefore, it may be advantageous in the long term to contain
contaminants originating from small-scale events such as RDDs before they infiltrate sewage and
drainage systems. These countermeasures would be most favorable in urban areas within days of
initial deposition or before a major rain/water washdown event.

6.3. Resuspension

Resuspension is a viable cause for concern arising from radiological incidents, and is the primary
mechanism responsible for increased respirable dose and for the lateral expanse of contamination
following the initial plume.  This is especially evident within the first few weeks of initial
deposition. Resuspension may be initiated by either natural or human-derived mechanisms, the
latter being more efficient. Accordingly, human activities are closely associated with urban areas
where resuspension is most prominent. Resuspension is a primary concern for two reasons: 1) as
mentioned above, the deposition area associated with RDDs is very limited, so the effects of
resuspension have a greater chance of increasing the expanse of contamination; and 2) RDDs are
likely to target populated areas with no warning, resulting  in mass panic and uncontrolled
evacuations. These actions will likely result in an abundance of human-aided resuspension, with
pedestrian and vehicle traffic being the largest contributors. Containment options should be
available for responders during evacuation and remediation. In spite of adverse effects associated
with resuspension, very little is understood in terms of potential, especially within urban areas.

6.4. Modeling and Simulation

Radiological models are a product of event-driven research, in that resources tend to fluctuate
and are contingent on the prevalence of large-scale events (e.g., Chernobyl). Resources for
research quickly surge immediately after an incident, then slowly dwindle as public awareness
wanes. Following Chernobyl, advancements in computer code enabled more sophisticated
modeling capabilities. A majority of these models are geared toward large-scale incidents,
though, or are strictly based on post-Chernobyl data. As highlighted by this paper, the
radiological threats to urban areas have evolved from large- (e.g., NWs) to small-scale threats
(e.g., RDDs). These smaller-scale threats present unique challenges that require advances in
current modeling practices and capabilities. Radiological models must account for the increased
spatial scales to include particle transport while also taking urban geometry into consideration.
Based on a literature review of radiological transport models, the following observations were

   •   Developers are starting to leverage  existing software platforms in  order to conserve
       resources that would otherwise be spent on developing pre-existing technologies (e.g.,
       urban planning software). This type of practice encourages developers and researchers
       alike to address specific gaps without having to start from scratch.
   •   Smaller event-driven models (i.e., multi-compartment models) that address  specific
       phases of a radiological emergency that would otherwise be too complex or resource-
       intensive to develop are becoming more prevalent. These compartment models can then
       be incorporated into larger platforms for addressing a range of radiological incidents.
   •   Different phases of a radiological incident may require stringent computer codes that are
       not  necessarily all-inclusive, which lends itself to the above observation [9]. Furthermore,
       as highlighted by this paper, there are a significant number of gaps in the literature
       regarding particle transport. These incompatibilities and research gaps result in
       increasingly complex models that require a large number of inputs in order to mitigate
       these uncertainties.

6.5. Conclusion

An extensive survey of the literature was conducted for the purpose of: 1) determining whether
empirical evidence exists for further characterizing RDDs according to NW detonation and NPP
accident literature, 2) providing an  overview and  analysis of the current state of literature related
to radiological sources with reference to particle transport, 3) contrasting  the behaviors of
radionuclides in urban and rural environments, and 4) exploring the current state of radiological
models, taking the above findings into consideration. Firstly, while the radiological sources
investigated by this paper (i.e., RDDs, NPP accidents, and NW detonations) differ in  magnitude
and overall impact on the environment, these disparities were more evident when NPP accidents
and NW detonations were compared to RDDs.  The differences were largely a function of the
impacted area, in that particle transport appears to be a localized problem. Therefore,  factoring in
the magnitude of NPP accidents and NW detonations, RDDs offer the greatest potential for the


dispersion of contamination. Second, explosive RDDs can be made of an array of materials,
which in turn influence the physiochemical properties of the dispersed contaminants.
Accordingly, greater uncertainty exists about RDDs' particle characteristics and transport. Third,
the environment in which the contaminant is deposited, and the spreading mechanisms that
environment hosts, greatly influence the behavior of the particle and the availability of exposure
pathways. Urban environments host an array of spreading mechanisms not present in rural areas.
The effects are augmented by the presence of less permeable surfaces that make up urban
environments, further encouraging the spreading of contaminants. Lastly, taking the above points
into consideration, more advanced models are needed to account for these physiochemical and
spatial variations. Current radiological models are adequate for large-scale incidents; however,
they lack spatial resolution to account for the inconsistencies afforded by RDDs, especially in
urban environments.

The findings of this literature review support a need for research in the transport of radionuclides
following a radiological incident in addition to more stringent computer codes capable of site
specific modeling. Site specific modeling is key to understanding the interaction of radiological
contaminants with the environment and how those interactions might affect potential radiation
exposure, remediation, and waste management strategies [35].  Site specific modeling results
would likely help determine potential contaminated hot spots, monitoring and sampling
locations, gross mitigation methods, and waste storage locations and techniques [35]. Based on
these findings, this report recommends future work as described in the subsections below.

       6.5.1.  Spreading Mechanisms
The migration of radionuclides in urban areas is poorly understood. Particularly, the interaction
and transport of radioactive contaminants within urban areas while accounting for infrastructure,
biomass, wastewater collection/treatment systems, and eventual disposal processes are of great
concern. Further research is needed to better understand the underlying processes and the overall
impacts of the following areas:

   •   The interaction (e.g., removal, and sorption) of radionuclides with surfaces and
       geometries commonly found in urban areas. This research might provide a better
       understanding of the dynamics that may lead to the migration of deposited particles,
       produced by an ROD, in urban areas.
   •   The magnitude and effects of anthropological resuspension in urban areas. Much is
       known in terms to resuspension in rural arid environments; although, very little data
       exists for sub-tropical urban environments. Literature suggests that contaminants residing
       in urban areas are more prone to resuspension. However, the impact of resuspension in
       urban areas is largely unknown. Such research might lend to improved modeling
       capabilities and, to a much greater extent, a better understanding of the implications of
       resuspension in urban areas.
   •   Although this report (for brevity's sake) does not discuss this subject in great detail, the
       literature lends itself to the need for near-term countermeasures that could reduce the
       spread of contamination, by means of human and natural activities, immediately
       following a radiological incident [25]. This paper would recommend that future research
       focus on fast-acting technologies  capable of affixing contaminants (i.e., retarding their


       movement) until more permanent options become available. This would potentially
       restrict the expanse of the contamination while emergency response and evacuation
       operations take place, which (the authors conclude) is of great importance immediately
       following initial deposition.
       6.5.2.  Modeling
Following previous radiological incidents, numerous prediction models have been developed to
estimate radiation exposure and to predict fate of radioactive materials in urban and rural areas.
Based on a high-level review assessing the current state of radiological models, this paper
recommends the following:

   •   Very little is known in terms of the movement and magnitude of water-soluble
       contaminants. It is suspected that these contaminants will follow the topography of the
       land and, depending on the extent of urbanization, eventually enter storm drain systems.
       Therefore, it may be advantageous to explore the use of commercial software
       applications for modeling urban runoff to better predict the migration of water-soluble
       radioactive contaminants in urban environments introduced via water applications (i.e.,
       decontamination activities, etc.) and precipitation (i.e., rain, snow, fog, etc.).

   •   Studies have shown that surface waters play an important role in particle transport,
       particularly in river banks and flood zones within close proximity to large rivers [35].
       Therefore, leveraging flood and  surface runoff models might be a viable option for
       predicting the impact of surface  water aided particle transport for rural areas.
In order to achieve the above recommendations, this report recommends the following actions:

   •   A more thorough technical review evaluating available models relevant to radiological

   •   Leveraging of existing software platforms by integrating radiological capabilities into
       pre-existing computer codes.

   •   The modularization of smaller event-based models that address specific gaps.

   •   The incorporation of Fukushima-derived measurements into radiological models.

   •   The development of an operational tool capable of supporting multi-compartment models
       for predicting the transport of radioactive materials within urban areas to support
       monitoring, remediation, and waste management decisions.


1.     Mikazu Yui, Overview ofJAEA 's R&Dfor Environmental Restoration in Fukushima and
      International Collaboration, J.A.E.A.J. Fukushima Environmental Safety Center, Editor.
      2015, US-Japan Decommissioning and Environmental Management WG: Japan, p. 17.
2.     P. Andrew Karam, Radiological Terrorism: Human And Ecological Risk Assessment, An
      InternationalJournal. Taylor & Francis, 2005. 11(3): p. 501-523.
3.     J. Peterson, et d.., Radiological and Chemical Fact Sheets to Support Health Risk
      Analyses for Contaminated Areas. 2007.
4.     S. Glasstone, P.D., The Effects of Nuclear Weapons. 1977: Washington, DC.
5.     Takeshi Fuj iwara, et al., Isotopic Ratio And Vertical Distribution of Radionuclides In Soil
      Affected By The Accident of Fukushima Dai-Ichi Nuclear Power Plants, ournal of
      Environmental Radioactivity 113 (2012) 37, 2012.  113: p. 37-44.
6.     B. Salbu, O.C. Lind, and L. Skipperud, Radionuclide Speciation And Its Relevance In
      Environmental Impact Assessments. Journal of Environmental Radioactivity 2004. 74: p.
7.     Admon, U., et al., Radioactive Particles In The Environment: Sources, Particle
      Characterization And Analytical Techniques. 2011, International Atomic Energy Agency,

8.     Kasper G. Andersson, et al., Revision Of Deposition And Weathering Parameters For
      The Ingestion Dose Module (Ecosys) Of The Argos AndRodos Decision Support Systems.
      Journal Of Environmental Radioactivity, 2011. 102: p. 1024-1031.
9.     J. Brown, et al., Requirements of Future Models For Inhabited Areas. Journal of
      Environmental Radioactivity, 2006. 85: p. 344-360.
10.    G. Rosner and R. Winkler, Long-Term Variation of Post-Chernobyl 90sr, 137cs, 238pu
      And 239,240pu Concentrations In Air, Depositions To Ground, Re suspension Factors
      AndResuspension Rates In South Germany. The Science of The Total Environment 2001.
      273: p. 11-25.
11.    G. Rosner, H. Hiitzl, and R. Winkler, Long-Term Behaviour of Plutonium In Air And
      Deposition And The Role of Resuspension In A Semi-Rural Environment In Germany.
      The Science of The Total Environment 1997. 196: p. 255-261.
12.    Mikhail Balonov, The Chernobyl Accident As A Source Of New  Radiological Knowledge:
      Implications For Fukushima Rehabilitation And Research Programmes. IOP Science,
      Journal of Radiological Protection, 2013. 33: p. 27-40.
13.    Frederick T. Harper, Stephen V. Musolino, and William B. Wente, Realistic Radiological
      Dispersal Device Hazard Boundaries And Ramifications For Early Consequence
      Management Decisions. Health Physics., 2007. 93(1): p. 1-16.
14.    Y. Hatano, et al., Aerosol Migration Near Chernobyl: Long-Term Data And Modeling.
      Atmospheric Environment, 1998. 32(14/15): p. 2587-2594.
15.    Mukhambetkali Burkitbayev, et al., Ecological Impacts Of Large-Scale War
      Preparations: Semipalatinsk Test Site, Kazakhstan. Springer Science+Business Media
      B.V. 2011,2011.
16.    Past And Recent Trends In Radioecology.  Environment International, 1994. 20(5): p.
17.    C. Duefias, et al., Long-Term Variation (1992-1999) of Gross-Beta, 210pb And 90sr
      Concentrations In Rainwater And Deposition To Ground. Journal of Geophysical
      Research. 108(D11): p. 4336.


18.     J. Christodouleas, R.F., C. Ainsley, Z. Tochner, S. Hahn, E. Glatstein,, Short-Term and
       Long-Term Health Risks of Nuclear-Power-Plant Accidents. N Engl J Med, 2011.
19.     Ole Christian Lind, et al., Overview Of Sources Of Radioactive Particles Of Nordic
       Relevance As Well As A Short Description Of Available Particle Characterization
       Techniques. 2008.
20.     Kownacka, L. Natural and artificial radionuclides in the tropospheric and lower
       stratospheric air over Poland, in International Meeting on Low-Level Air Radioactivity
       Monitoring. 2000. Nidzica, Poland.
21.     H. Mala, et al., Particle size distribution of radioactive aerosols after the Fukushima and
       the Chernobyl accidents. Journal of Environmental Radioactivity, 2013. 126: p. 92-98.
22.     K.G. Anders son, et al., Requirements For Estimation of Doses From Contaminants
       Dispersed By A 'Dirty Bomb' Explosion In An Urban Area. Journal of Environmental
       Radioactivity, 2009. 100: p. 1005-1011.
23.     Sang Don Lee, et al., Radiological Dispersal Device Outdoor Simulation Test: Cesium
       Chloride Particle Characteristics. Journal of Hazardous Materials 2010. 176: p. 56-63.
24.     Vladimir P.  Reshetin, Estimation of Radioactivity Levels Associated With A 90sr Dirty
       Bomb Event. Atmospheric Environment, 2005. 39: p. 4471-4477.
25.     K. Andersson, Airborne Radioactive Contamination in Inhabited Areas. Radioactivity in
       the Environment, ed. K Andersson. 2009:  Elsevier Science.
26.     M. L. Wesely and B. B. Hicks, A Review Of The Current Status Of Knowledge On Dry
       Deposition.  Atmospheric Environment 2000. 34: p. 2261-2282.
27.     K.G. Andersson and J. Roed, Estimation of Doses Received In A Dry-Contaminated
       Residential Area In The Bryansk Region, Russia, Since The Chernobyl Accident. Journal
       of Environmental Radioactivity, 2006. 85: p. 228-240.
28.     Tone D. Bergan, Radioactive Fallout In Norway From Atmospheric Nuclear Weapons
       Tests. Journal of Environmental Radioactivity, 2002. 60: p. 189-208.
29.     B. Erlandsson and M. Isaksson, Relation Between The Air Activity And The Deposition of
       Chernobyl Debris. Environment International,, 1988. 14: p. 165-175.
30.     V. Ramzaev, et al., Radiocesium Fallout In The Grasslands On Sakhalin, Kunashir And
       Shikotan Islands Due To Fukushima Accident: The Radioactive Contamination of Soil
       And Plants In 2011.  Journal of Environmental Radioactivity, 2013. 118: p. 128-142.
31.     Gael Le Roux, et al., Deposition of Artificial Radionuclides From Atmospheric Nuclear
       Weapon Tests Estimated By Soil Inventories In French Areas Low-Impacted By
       Chernobyl. Journal of Environmental Radioactivity 2010. 101: p. 211-218.
32.     Brant Ulsh,  Steven Rademacher, and F. Ward  Whicker, Variations of 137cs Depositions
       And Soil Concentrations Between Alpine And Montane Soils In Northern Colorado.
       Journal of Environmental Radioactivity 2000.  47: p. 57-70.
33.     Lage Jonsson, et al., Various Consequences Regarding Hypothetical Dispersion of
       Airborne Radioactivity In A City Center. Journal of Environmental Radioactivity, 2013.
       116: p. 99-113.
34.     K. G. Andersson, J. Roed, and C. L. Fogh, Weathering Of Radiocaesium Contamination
       On Urban Streets, Walls And Roofs. Environmental Radioactivity, 2002. 62: p. 49-60.
35.     Center, F.E.S., Remediation of Contaminated Areas in the Aftermath of the Accident at
       the Fukushima Daiichi Nuclear Power Station: Overview, Analysis and Lessons Learned
       Part 2: Recent Developments, Supporting R&D and International Discussions. 2015,
       Japan Atomic Energy Agency.


36.     Andrian A. Seleznev, Ilia V. Yarmoshenko, and Alexey A. Ekidin, Accumulation of
       137cs In Puddle Sediments Within Urban Ecosystem. Journal of Environmental
       Radioactivity, 2010. 101: p. 643-646.
37.     Eduardo B. Farfan, et al., Assessment Of Beta Particle Flux From Surface Contamination
       As A Relative Indicator For Radionuclide Distribution On External Surfaces Of A
       Multistory Building In Pripyat. Health Physics Journal, 2010.
38.     Roed, J., K. G. Andersson, and O. Togawa. Weathering of radionuclides deposited in
       inhabited areas, in 1996 International congress on radiation protection. Proceedings.
       1996. Seibersdorf: IRPA, 1996. .
39.     Konstantin Volcheka, et  al., Adsorption Of Cesium On Cement Mortar From Aqueous
       Solutions. Journal Of Hazardous Materials 2011. 194: p. 331-337.
40.     Nao Kamei-Ishikawa, et  al., Fate ofRadiocesium In Sewage Treatment Process Released
       By The Nuclear Accident At Fukushima. Chemosphere 2013. 93: p. 689-694.
41.     Nao Kamei-Ishikawa, Ayumi Ito, and Teruyuki Umita, Fate of Stable Strontium In The
       Sewage Treatment Process As An Analog For Radiostrontium Released By Nuclear
       Accidents. Journal of Hazardous Materials, 2013. 260: p. 420-424.
42.     Simon V.  Avery, Fate of Caesium In The Environment: Distribution Between The Abiotic
       AndBiotic Components of Aquatic And Terrestrial Ecosystems. Journal of Environmental
       Radioactivity, 1996. 30(2): p. 139-171.
43.     Eduardo Gallego, Mud: A Model To Investigate The Migration of 13 7cs In The Urban
       Environment And Drainage And Sew age Treatment Systems. Journal of Environmental
       Radioactivity, 2006. 85:  p.  247-264.
44.     F. A.  Tikhomirov, A. I. Shcheglov, and V. P. Sidorov, Forests And Forestry: Radiation
       With Special Reference To  The Zone Protection Chernobyl Measures Accident.  The
       Science of The Total Environment, 1993. 137: p. 289-305.
45.     Saito, K.,  Mapping and Modelling of Radionuclide Distribution on the Ground due to the
       Fukushima Accident. Radiation Protection Dosimetry, 2014. 160(4): p. 283-287.
46.     A. D. Horrill, et al., The Effect of Heather Burning On The Transfer ofRadiocaesium To
       Smoke And The Solubility ofRadiocaesium Associated With Different Types of Heather
       Ash. Journal of Environmental Radioactivity, 1995. 29(1): p. 1-10.
47.     G. Shaw, Radioactivity in the Terrestrial Environment. Vol. 10. 2007: Elsevier.
48.     P. J. P. Bonnett and M. A. Anderson, Radiocaesium Dynamics In A Coniferous Forest
       Canopy: A Mid-Wales Case Study. The Science of The Total Environment, 1993. 136:  p.
49.     T. M. Nakanishi, N. I. Kobayashi, and K. Tanoi,  Radioactive Cesium Deposition On
       Rice,  Wheat, Peach Tree And Soil After Nuclear Accident In Fukushima. Journal of
       Radioanalytical and Nuclear Chemistry 2013. 296: p. 985-989.
50.     Christian  Lange Fogh and Kasper Grann Andersson, Dynamic Behaviour of 137cs
       Contamination In Trees of The Briansk Region, Russia.  The Science of the Total
       Environment, 2001. 269: p. 105-115.
51.     PekkaNygren, et al., Behaviour of 137cs From Chernobyl Fallout In A Scots Pine
       Canopy In Southern Finland. Canadian Journal of Forest Research, 1994. 24: p. 1210-

52.    K. lijima, et al. Distribution of Radioactive Cesium in Trees and Effect of
      Decontamination of Forest Contaminated by the Fukushima Nuclear Accident, in
      Proceedings of the ASME 2013 15th International Conference on Environmental
      Remediation and Radioactive Waste Management. 2013. Brussels, Belgium.
53.    Takahiro Nakanishi, etal., 137cs Vertical Migration In A Deciduous Forest Soil
      Following The Fukushima Dai-Ichi Nuclear Power Plant Accident. 2014. 128.
54.    Gabriele Clooth and D. C. Aurnann, Environmental Transfer Parameters And
      Radiological Impact of The Chernobyl Fallout In And Around Bonn (FRG). Journal of
      Environmental Radioactivity, 1990. 12: p. 97-119.
55.    D. Bugai, et al., Characterization Of Subsurface Geometry And Radioactivity
      Distribution In The Trench Containing Chernobyl Clean-Up Wastes. Environmental
      Geology (aka Environmental Earth Sciences), 2005. 47: p. 869-881.
56.    Katsushi Kuroda, Akira Kagawa, and Mario Tonosaki, Radiocesium Concentrations In
      The Bark, Sapwood And Heartwood of Three Tree Species Collected At Fukushima
      Forests Half A Year After The Fukushima Dai-Ichi Nuclear Accident. Journal of
      Environmental Radioactivity, 2013. 122: p. 37-42.
57.    K. Bunzl, et al., Spatial Variability Of The Vertical Migration Of Fallout 137cs In The
      Soil Of A Pasture, And Consequences For Long-Term Predictions. Radiation and
      Environmental Biophysics, 2000. 39: p. 197-205.
58.    Hiroaki Kato, Yuichi Onda, and Mengistu Teramage, Depth Distribution of 137cs,  134cs,
      And 1311 In Soil Profile After Fukushima Dai-Ichi Nuclear Power Plant Accident.
      Journal of Environmental Radioactivity, 2012. Ill: p. 59-64.
59.    S. Almgren and M. Isaksson, Vertical Migration Studies Of 137cs From Nuclear
      Weapons Fallout And The Chernobyl Accident. Journal Of Environmental Radioactivity,
      2006. 91: p. 90-102.
60.    C. Duffa and P. Renaud, 238pu And 239+240pu Inventory And Distribution Through The
      Lower Rhone Valley Terrestrial Environment (Southern France). Science of The Total
      Environment, 2005. 348: p. 164-172.
61.    J. Solecki and S. Chibowski, Studies on Horizontal and Vertical Migration of90Sr in Soil
      Systems. Polish Journal of Environmental Studies, 2002. 11(2): p. 157-164.
62.    F. R. Livens and M. S. Baxter, Particle Size AndRadionuclide Levels In Some West Soils
      Cumbrian. The Science of The Total Environment, 1988. 70: p. 1-17.
63.    A. Baezaa, et al., Seasonal Variations In Radionuclide Transfer In A Mediterranean
      Grazing-LandEcosystem. Journal of Environmental Radioactivity, 2001. 55: p. 283-302.
64.    Rob N. J. Comans, et al., Mobilization of Radiocaesium In Pore Water of Lake
      Sediments. Nature Publishing Group,  1989. 339.
65.    Tomoko Ohta, et al., Prediction ofGroundwater Contamination With 137cs And 1311
      From The Fukushima Nuclear Accident In The Kanto District. Journal of Environmental
      Radioactivity, 2012. Ill: p. 38-41.
66.    Luigi Monte, et al., Assessment Of State-Of-The-Art Models For Predicting The
      Remobilisation of Radionuclides Following The Flooding of Heavily Contaminated
      Areas: The Case Of Pripyat River Floodplain. Journal Of Environmental Radioactivity,
      2006. 88: p. 267-288.

67.     Scott C James, Craig Jones, and Jesse D. Roberts, Consequence Management, Recovery
       And Restoration After A Contamination Event. 2005, U. S. Department Of Energy, Office
       of Scientific and Technical Information: Albuqueque, New Mexico.
68.     Shinji Ueda, et al., Fluvial Discharges ofRadiocaesium From Watersheds Contaminated
       By The Fukushima Dai-Ichi Nuclear Power Plant Accident, Japan. Journal of
       Environmental Radioactivity, 2013. 118: p. 96-104.
69.     Y. Onishi, et al., Aquatic Assessment Of The Chernobyl Nuclear Accident And Its
       Remediation. American Society of Civil Engineers, 2007. 133(11): p. 1015-1023.
70.     G. A. Sehmel, Particle Resuspension: A Review. Environment International, 1980. 4: p.
71.     K. M. Thiessen, et al., Modelling Radionuclide Distribution And Transport In The
       Environment. Environmental Pollution 1999. 100: p. 151-177.
72.     K. W. Nicholson, Review Article: A Review Of Particle Resuspension. Atmospheric
       Environmental, 1988. 22(12): p. 2639-2651.
73.     David C. Kocher,  John R. Trabalka, and A. lulian Apostoaei, Derivation Of Effective
       Resuspension Factors In Scenarios For Inhalation Exposure Involving Resuspension Of
       Previously Deposited Fallout By Nuclear Detonations At Nevada Test Site, I. Senes Oak
       Ridge, Editor. 2009, U.S. Department Of Defense, Defense Threat Reduction Agency.
74.     K. W. Nicholson and J. R. Branson, Factors Affecting Resuspension By Road Traffic. The
       Science of The Total Environment, 1990. 93: p. 349-358.
75.     NKA proj ect AKTU-200, Environmental Consequences of Releases From Nuclear
       Accidents A Nordic Perspective, U.I.F.E. Technology, Editor. 1990, Nordic Liaison
       Committee For Atomic Energy.
76.     J. A. Garland and  I. R. Pomeroy, Resuspension of Fall-Out Material Following The
       Chernobyl Accident. Journal of Aerosol Science 1994. 25(5): p. 793-806.
77.     T. G. Hinton, Contamination of Plants By Resuspension: A Review, With Critique of
       Measurement Methods. The Science of The Total Environment, 1992. 121: p. 177-193.
78.     J. A. Garland, et al., Modelling Of Resuspension, Seasonally And Losses During Food
       Processing First Report Of The Vamp Terrestrial Working Group. 1992, International
       Atomic Energy Agency,  The Vamp Terrestrial Working Group: laea Vienna.
79.     Zitouni Ould-Dada and Nasser M. Baghini, Resuspension of Small Particles From Tree
       Surfaces. Atmospheric Environment, 2001. 35: p. 3799-3809.
80.     C. Walsh, Calculation Of Resuspension Doses For Emergency Response. 2002,
       Environmental Assessments Department and Emergency Response Group Quality
       Management System.
81.     A. lulian Apostoaei and David C. Kocher, Radiation Doses To Skin From Dermal
       Contamination. 2010, U.S. Department Of Defense, Defense Threat Reduction Agency,.
82.     V. A. Kashparov,  et al., Resuspension of Radionuclides And The Contamination of
       Village Areas Around Chernobyl. Journal of Aerosol Science 1994. 25(5): p. 755-759.
83.     Hiroko Kidoa, et al., The Simulation of Long-Range Transport of 137cs From East Asia
       To Japan In 2002  And 2006. Journal of Environmental Radioactivity 2012. 103: p. 7-14.
84.     Mai K. Pham, et al., Dry And Wet Deposition of7be, 2lOpb And 137cs In Monaco Air
       During 1998e2010: Seasonal Variations of Deposition Fluxes. Journal of Environmental
       Radioactivity, 2013. 120: p. 45-57.
85.     Werner Hollander, Resuspension Factors of 137cs In Hannover After The Chernobyl
       Accident. Journal of Aerosol Science 1994. 25(5): p. 789-792.


86.     Garland, J. A. Surface Deposition From Radioactive Plumes, in Proceedings of the
       seminar on radioactive releases and their dispersion in the atmosphere following a
       hypothetical reactor accident. 1980. UKAEA Atomic Energy Research Establishment,
       Harwell: Risco.
87.     J. W. Healy. An Overview Of Re suspension Models: Application To Low Level Waste
       Management. 1980. Los Alamos Scientific Laboratory.
88.     E. K. Garger, Air Concentrations Of Radionuclides In The Vicinity Of Chernobyl And
       The Effects Of Re suspension. J. Aerosol Sci.,, 1994. 25(5): p. 745-753.
89.     J. A. Garland, Resuspension Of Particulate Matter From Grass And Soil. 1979,
       Environmental & Medical Sciences Division, p. 35.
90.     V. A. Kashparov, et al., Forest Fires In The Territory Contaminated As A Result of The
       Chernobyl Accident: Radioactive Aerosol Resuspension And Exposure of Fire-fighters.
       Journal of Environmental Radioactivity 2000. 51: p. 281-298.
91.     O. Guillitte and C. Willdrodt, An Assessment Of Experimental And Potential
       Countermeasures To Reduce Radionuclide Transfers In Forest Ecosystems. The Science
       Of The Total Environment, 1993. 137: p. 273-288.
92.     V. I. Yoschenko, et al., Resuspension And Redistribution of Radionuclides During
       Grassland And Forest Fires In The Chernobyl Exclusion Zone: Part I. Fire Experiments.
       Journal of Environmental Radioactivity, 2006. 86: p. 143-163.
93.     TA. Mousseau, et al., Highly reduced mass loss rates and increased litter layer in
       radioactively contaminated areas. Oecologia, 2014. 175: p. 429-37.
94.     K. W. Nicholson, et al., The Effects of Vehicle Activity On Particle Resuspension. Journal
       of Aerosol Science, 1989. 20(8): p. 1425-1428.
95.     G. A. Jsmel, Particle Resuspension From An Asphalt Road Caused By Car And Truck
       Traffic. Atmospheric Environment, 1973. 7: p. 291-309.
96.     G. A. Sehmel, Transuranic And Tracer Simulant Resuspension. 1977, The Energy
       Research And Development Administration.
97.     Joseph H. Shinn, Donald N. Homan, and William L. Robison, From Restoration Of
       Environments Affected By Residues From Radiological Accidents: Approaches To
       Decision Making., in Resuspension Studies At Kikini Atoll. 1980, U. S. Department Of
       Energy, Office Of Scientific And Technical Information.
98.     G. S. Linsley, Resuspension Of The Transuranium Elements-A Review Of Existing Data.
       1978, National Radiological Protection Board: Harwell, Didcot, Oxfordshire, England, p.
99.     N. Yamaguchi, et al., Radiocesium And Radioiodine In Soil Particles Agitated By
       Agricultural Practices: Field Observation After The Fukushima Nuclear Accident.
       Science of The Total Environment 2012. 425: p. 128-134.
100.   Tanaka, A. Indoor Contamination from the Fukushima Nuclear Power Plant Incident, in
       2075 EPA International Decontamination Research and Development Conference. 2015.
       Durham, NC.
101.   Jerzy Bartnicki, et al. Atmospheric Transport and Deposition of Radioactive Particles
      from Potential Accidents at Kola Nuclear Power Plant, in 1st Joint Emergency.
       Preparation.  & Response/Robotic & Remote System. Top. Meeting. 2006. Salt Lake City,
102.   M.A. Pires Rio and H.G. Paretzke, The Spread Of 137 CS Resuspension Of Contaminated
       Soil In The Urban Area Goidnia. 2000. p. 249-257.


103.   G. A. Loosmore, Evaluation and Development of Models for Resuspension of Aerosols at
      Short Times after Deposition. 2002, U.S. Department of Energy, Lawrence Livermore
      National Laboratory, Atmospheric Sciences Division.
104.   K. W. Nicholson, Wind Tunnel Experiments On The Resuspension of Paniculate
      Material Atmospheric Environment 1993. 27a(2): p. 181-188.
105.   C. Papastefanou, et al., Atmospheric Deposition ofCosmogenic 7be And 137cfsr Om
      Fallout of The Chernobyl Accident. The Science of The Total Environment 1995. 170: p.
106.   Programme, E.M.f.R.S.E., Environmental Modelling of Remediation of Urban
      Contaminated Areas Report of the Urban Remediation Working Group ofEMRAS Theme
      2 2007, The International Atomic Energy Agency, p. 509.
107.   EPA, U.S.,  WARRP Decon-13: Subject Matter Expert (SME) Meeting Waste Screening
      and Waste Minimization Methodologies Project. 2013.
108.   Goel, M.K., Runoff Coefficient. Encyclopedia of Snow, Ice and Glaciers, 2014: p. 952-

 United States
 Environmental Protection
   PERM IT NO. G-35
Office of Research and Development (8101R)
Washington, DC 20460

 Official Business
 Penalty for Private Use