Technical Reports SeriEs No.

424

Remediation of Sites with
Dispersed Radioactive
Contamination

REMEDIATION OF SITES WITHDISPERSED RADIOACTIVECONTAMINATIONThe following States are Members of the International Atomic Energy Agency:
AFGHANISTANALBANIA
ALGERIAANGOLA
ARGENTINA
ARMENIA
AUSTRALIA
AUSTRIAAZERBAIJANBANGLADESH
BELARUSBELGIUM
BENIN
BOLIVIABOSNIA AND HERZEGOVINABOTSWANABRAZIL
BULGARIA
BURKINA FASOCAMEROONCANADACENTRAL AFRICAN
REPUBLIC
CHILE
CHINA
COLOMBIACOSTA RICA
CÔTE D’IVOIRECROATIACUBACYPRUSCZECH REPUBLIC
DEMOCRATIC REPUBLIC
OF THE CONGO
DENMARKDOMINICAN REPUBLICECUADOREGYPTEL SALVADORERITREAESTONIAETHIOPIAFINLANDFRANCEGABONGEORGIAGERMANYGHANA
GREECEGUATEMALAHAITI
HOLY SEE
HONDURASHUNGARYICELAND
INDIA
INDONESIA
IRAN, ISLAMIC REPUBLIC OFIRAQIRELANDISRAELITALY
JAMAICAJAPANJORDANKAZAKHSTANKENYAKOREA, REPUBLIC OFKUWAITKYRGYZSTANLATVIALEBANON
LIBERIA
LIBYAN ARAB JAMAHIRIYALIECHTENSTEIN
LITHUANIALUXEMBOURG
MADAGASCARMALAYSIAMALI
MALTA
MARSHALL ISLANDS
MAURITIUS
MEXICOMONACOMONGOLIA
MOROCCOMYANMAR
NAMIBIANETHERLANDSNEW ZEALANDNICARAGUANIGERNIGERIANORWAYPAKISTAN
PANAMAPARAGUAYPERU
PHILIPPINESPOLANDPORTUGALQATARREPUBLIC OF MOLDOVAROMANIARUSSIAN FEDERATIONSAUDI ARABIASENEGALSERBIA AND MONTENEGROSEYCHELLESSIERRA LEONESINGAPORESLOVAKIASLOVENIA
SOUTH AFRICASPAIN
SRI LANKA
SUDAN
SWEDENSWITZERLANDSYRIAN ARAB REPUBLICTAJIKISTANTHAILANDTHE FORMER YUGOSLAVREPUBLIC OF MACEDONIATUNISIA
TURKEYUGANDAUKRAINE
UNITED ARAB EMIRATESUNITED KINGDOM OFGREAT BRITAIN ANDNORTHERN IRELAND
UNITED REPUBLICOF TANZANIAUNITED STATES OF AMERICAURUGUAYUZBEKISTANVENEZUELAVIETNAMYEMENZAMBIA
ZIMBABWE

The Agency’s Statute was approved on 23 October 1956 by the Conference on the Statute of
the IAEA held at United Nations Headquarters, New York; it entered into force on 29 July 1957.
The Headquarters of the Agency are situated in Vienna. Its principal objective is “to accelerate andenlarge the contribution of atomic energy to peace, health and prosperity throughout the world’’.
© IAEA, 2004
Permission to reproduce or translate the information contained in this publication may beobtained by writing to the International Atomic Energy Agency, Wagramer Strasse 5, P.O. Box 100,
A-1400 Vienna, Austria.
Printed by the IAEA in Austria
June 2004
STI/DOC/010/424

TECHNICAL REPORTS SERIES No. 424

REMEDIATION OF SITES WITHDISPERSED RADIOACTIVECONTAMINATIONINTERNATIONAL ATOMIC ENERGY AGENCY
VIENNA, 2004

IAEA Library Cataloguing in Publication Data
Remediation of sites with dispersed radioactive contamination. —Vienna : International Atomic Energy Agency, 2004.
p. ; 24 cm. — (Technical reports series, ISSN 0074–1914 ; no. 424)
STI/DOC/010/424
ISBN 92–0–114603–5
Includes bibliographical references.
1. Radioactive decontamination. 2. Hazardous waste site remediation.
I. International Atomic Energy Agency. II. Technical reports series(International Atomic Energy Agency) ; 424.
IAEAL

04–00357

FOREWORD
The IAEA attaches great importance to the dissemination of informationthat can assist Member States with the development, implementation,
maintenance and continuous improvement of systems, programmes andactivities that support a sustainable nuclear fuel cycle and nuclear applications,
including the legacy of past practices and accidents. The IAEA has therefore
initiated a comprehensive programme of work covering all aspects ofenvironmental remediation:
(a)
(b)
(c)
(d)
(e)
(f)

Technical and non-technical factors influencing a decision on
environmental remediation;
Site characterization techniques and strategies;
The assessment of remediation techniques;
The assessment of technical options for the cleanup of contaminatedmedia;
Post-restoration compliance monitoring;
The assessment of the costs of remediation measures.

Dispersed low level contamination poses a particular challenge to thosecharged with its remediation. Many techniques are not efficient below certain
concentration thresholds or entail more severe impacts on certain
environmental compartments than the contamination itself. In such cases
justification for remediation may not be given on radiation protection grounds,
but remediation may still be demanded by the public.
This report examines a variety of technological options for dealing withdispersed low level contamination. The approaches are broadly grouped into
the three categories of:
(1)
(2)
(3)

Non-intervention;
Containment;
Removal.

The techniques related to the above three categories may need to bechosen in conjunction or consecutively for any given site in order to arrive at an
efficient and long term solution. These techniques are briefly outlined and theiradvantages and limitations are discussed. The need for a holistic design for the
remedial action is stressed.
The extensive body of references provided will help the reader to findmore detailed information.

The IAEA officer responsible for the preparation of this report was W.E.
Falck of the Division of Nuclear Fuel Cycle and Waste Management.

EDITORIAL NOTE
The use of particular designations of countries or territories does not imply any
judgement by the publisher, the IAEA, as to the legal status of such countries or territories,of their authorities and institutions or of the delimitation of their boundaries.
The mention of names of specific companies or products (whether or not indicatedas registered) does not imply any intention to infringe proprietary rights, nor should it beconstrued as an endorsement or recommenation on the part of the IAEA.

CONTENTS
1.

INTRODUCTION . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

1

1.1. Background . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
1.2. Scope . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
1.3. Structure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

1
6
6

2.

DISPERSED RADIOACTIVE CONTAMINATION . . . . . . . . . .

6

3.

FACTORS INFLUENCING TECHNOLOGY SELECTION . . . .

8

3.1.
3.2.
3.3.
3.4.
3.5.
3.6.
3.7.
3.8.

Fundamental technical choices . . . . . . . . . . . . . . . . . . . . . . . . . .
Technology sources . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Effectiveness . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Cost . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Occupational safety and health . . . . . . . . . . . . . . . . . . . . . . . . . .
Secondary environmental impacts . . . . . . . . . . . . . . . . . . . . . . . .
Prior experience . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Socioeconomic considerations . . . . . . . . . . . . . . . . . . . . . . . . . . .

8
10
11
12
12
13
13
14

NON-INTERVENTION . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

14

4.1. Basic considerations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
4.2. Dilution and dispersion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
4.3. Natural attenuation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
4.3.1. Principles . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
4.3.2. Physical processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
4.3.2.1. Radioactive decay . . . . . . . . . . . . . . . . . . . . . . .
4.3.2.2. Filtration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
4.3.2.3. Volatilization . . . . . . . . . . . . . . . . . . . . . . . . . . .
4.3.3. Chemical processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
4.3.3.1. Precipitation . . . . . . . . . . . . . . . . . . . . . . . . . . . .
4.3.3.2. Co-precipitation . . . . . . . . . . . . . . . . . . . . . . . .
4.3.3.3. Sorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
4.3.3.4. Complexation by organics . . . . . . . . . . . . . . . .
4.3.4. Biological processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
4.3.5. Definitions and applicability . . . . . . . . . . . . . . . . . . . . . .
4.4. Alternative land uses . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
4.5. Agricultural countermeasures . . . . . . . . . . . . . . . . . . . . . . . . . . .

14
15
15
15
16
16
16
16
18
18
18
19
19
20
20
21
26

4.

5.

6.

7.

CONTAINMENT . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

29

5.1.
5.2.
5.3.
5.4.
5.5.
5.6.
5.7.

Enhanced attenuation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Low permeability barriers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Permeable reactive barriers . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Immobilization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Biological barrier walls (biowalls) . . . . . . . . . . . . . . . . . . . . . . . .
Phytostabilization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Constructed wetlands . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

29
30
31
38
40
42
45

REMOVAL . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

49

6.1. Preconditions and constraints . . . . . . . . . . . . . . . . . . . . . . . . . . .
6.2. Immobilization and solidification (ex situ) . . . . . . . . . . . . . . . . .
6.3. Surface and groundwaters: pump and treat . . . . . . . . . . . . . . . .
6.4. Enhanced recovery . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
6.5. Chemical extraction (ex situ) . . . . . . . . . . . . . . . . . . . . . . . . . . . .
6.6. Hydrometallurgical methods . . . . . . . . . . . . . . . . . . . . . . . . . . . .
6.7. Segregation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
6.8. Biosorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
6.9. Bioleaching . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
6.10. Phytoextraction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
6.10.1. Overview . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
6.10.2. Uranium removal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
6.10.3. Strontium removal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
6.10.4. Caesium removal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
6.10.5. Phytoextraction project in Belarus . . . . . . . . . . . . . . . . .
6.11. Rhizofiltration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

49
49
50
52
54
55
56
56
59
63
63
67
68
69
71
72

SUMMARY AND CONCLUSIONS . . . . . . . . . . . . . . . . . . . . . . . . .

75

REFERENCES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 79
GLOSSARY . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 115
CONTRIBUTORS TO DRAFTING AND REVIEW . . . . . . . . . . . . . . . 117

1. INTRODUCTION
1.1. BACKGROUND
To respond to the needs of Member States, the IAEA launched anenvironmental remediation project to deal with the problems of radioactive
contamination worldwide.
The term ‘remediation’ is used in this report to encompass all activities
leading to reduced exposure to radiation and to an improved environmental
and/or economic value of a site; it does not necessarily imply restoration to apristine environmental state. The terms ‘rehabilitation’ and ‘restoration’ areoften used interchangeably. In the context of this report remediation is taken tomean management of contamination (i.e. removal, containment and monitored
non-intervention).
The IAEA environmental remediation project includes an IAEA Coordinated Research Project [1], as well as the participation of IAEA experts inconcrete remediation projects when requested by individual Member States.
The IAEA has prepared several documents dedicated to particulartechnical or conceptual areas (see Table 1), including documents on the characterization of contaminated sites [8, 14], technical and non-technical factors
relevant to the selection of a preferred remediation strategy and technique [3,17], overview of applicable techniques for environmental remediation [5],options for the cleanup of contaminated groundwater [10] and planning andmanagement issues [19, 20]. In addition, a number of other IAEA publications
dealing with related aspects have been compiled under different IAEA
projects; these include TECDOCs on the remediation of uranium mill tailings,
the decontamination of buildings and roads and the characterization of decommissioned sites.
Detailed procedures for the planning and implementation of remedial
measures have been developed over the past decade or so. A critical element is
the characterization of the contamination and of the various environmental
compartments in which it is found, in order to be able to evaluate the applicability of remediation techniques [8]. The chemical or mineralogical form of thecontaminant will critically influence the efficiency of the remediation techniquechosen. Careful delineation of the contamination will ensure that only thoseareas or volumes of material that are actually contaminated are treated [8].
This, in turn, reduces the amount of any secondary waste generated. Understanding the processes of migration and dispersal is particularly importantwhen natural attenuation is to be relied upon as a mechanism for remediationand dose prevention (see Section 4.1).

1

TABLE 1.
RELATIONSHIP BETWEEN RELEVANT TOPICAL
DOCUMENTS PRODUCED OR UNDER DEVELOPMENT
SafetyManagement

Management of Factors for
Formulating aRadioactive
Waste from the Strategy for
Environmental
Mining andMilling of Ores Restoration [3][2]Databases

Technology

A Directory of Technologies for Extent ofInformation
the Remediation Environmental
Resources onof Radioactively Contamination
Radioactive
Contaminated by Naturally
Waste
OccurringSites [5]Management,
Radioactive
Decontamination
Materialand(NORM) andTechnological
Decommissioning,
Options forand
Mitigation [6]Environmental
Restoration [4]Monitoring and
Surveillance ofResidues from
the Mining andMilling of
Uranium andThorium [7]Characterization Design Criteria Technicalof Radioactively for a Worldwide Options for the
Remediation ofContaminated Directory ofContaminatedRadioactively
Sites for
GroundwaterContaminated
Remediation
Sites (DRCS) [9] [10]Purposes [8]Remediation ofAreas
Contaminatedby PastActivities andAccidents [11]Compliance
Monitoring for
RemediatedSites [12]2Special topics

The Long-term
Stabilization of
Uranium Mill
Tailings: Final
Report on the
Co-ordinatedResearch
Project 2000–
2004 [1]Remediation of
Site
A Worldwide
Characterization Sites
Directory ofTechniques Used ContaminatedRadioactively
by Hazardous
inContaminatedEnvironmental and Radioactive
Sites (DRCS),
Substances [15]Restoration
http://wwwdrcs.iaea.org [13] Activities [14]TABLE 1.
RELATIONSHIP BETWEEN RELEVANT TOPICAL
DOCUMENTS PRODUCED OR UNDER DEVELOPMENT (cont.)
SafetyManagement

Non-technical
Derivation of
Factors
Remediation
Impacting on the
Values for
Decision Making
Areas with
Processes inRadioactive
Residues from Environmental
Past Activities Remediation [17]and Accidents
and their
Implementation[16]

Databases

Technology

Special topics

This report

Removal of Sites
and Buildings
from Regulatory
Control Upon
the Terminationof Practices [18]Although contamination may originally be dispersed, secondary
processes may concentrate, fractionate or otherwise redistribute the radionuclides; for example, it was observed that Chernobyl fallout was concentrated ingeomorphological depressions by near surface migration processes (Fig. 1) [21].
Conversely, erosion may lead to the dispersion of radionuclides, particularly inmining environments.
The application of a remediation technique requires holistic studies
examining the technical feasibility of the proposed measures, including
analyses of their impact. Consequently, input from various scientific andengineering disciplines, including the health sciences, chemistry, physics,
geology, microbiology and environmental engineering, is necessary in order todevelop technical solutions. It is also necessary to include information on thepolitical, social and economic context. Factors such as the overall project costversus availability of resources over time, public perception and the availabilityof skilled workers need to be considered [17].

3FIG. 1. Schematic representation of radionuclide accumulation in topographical depressions due to near surface migration, soil erosion, etc. [21].

It is essential that the process of decision making be transparent, as thismakes the decision taken more defensible to possible public scrutiny and is also
an essential element of quality control. Formal methods to aid decision making
have been developed over the years. A recent IAEA document describes thesein more detail [17]. A generic scheme of decision making and technology
assessment is given in Fig. 2.
Techniques for remediating sites with well defined contaminated areas
containing relatively high concentrations of contaminants are well established.
The remediation of dispersed contamination, however, still constitutes achallenge when considering factors such as expected dose, cost, public
perception and anxiety, and minimal disturbance of the environment. Any
remedial action in an area containing dispersed contamination has to bejustified, in that it must remove or reduce the source term or prevent thelikelihood of exposure to a harmful dose. In order to optimize the intervention,a reasonable balance has to be found between the risk of exposure due to thepotential radiation dose and the cost of the remedial action [22]. Similar considerations apply to further lowering the residual contamination over larger areas
following prior remedial measures.
As large scale remediation operations can be disruptive to agriculture,
natural ecosystems or (engineered) waste management sites, suitable implementation strategies that take into account the socioeconomic boundary
conditions need to be developed [20]. An earlier IAEA publication, which wasdeveloped following the Chernobyl accident, focused on the planning of
emergency response and remediation measures [20]. Since its publication in
1991, extensive experience in related fields has been gained [23]. One of themain conclusions from this experience is that the answer to the problem should

4

Environmental
requirements

Problem
identification

Safetyrequirements

Scoping sitecharacterization

Definitionof goals

Stakeholders
No

Detailed sitecharacterization

Achievable?
Yes

No
Sufficient?

Technology
identification

Yes

Stakeholders
No
Time

Feasible/safe?
Yes
Government
policyTechnology
selectionNo

If technology
proves unsuitable

Technology
implementation

Stakeholders

Successful?
Yes
Government
policyCompliancemonitoring

Stakeholders

FIG. 2. Phases of decision making in remediation strategies and technologies [17].

be sought in low cost, low intensity, low maintenance (‘passive’) and often low
technology solutions. Treating the contamination typically gave rise to the
problem of having to dispose of large quantities of slightly contaminated5material, mostly topsoil [24]. In the case of short lived nuclides, in situ methods
to prevent direct exposure or uptake via the food pathway have been successfully developed [25].

1.2. SCOPE
This report describes remediation techniques that are applicable todispersed radioactive contamination at a variety of sites, including surface soil,
the vadose zone, surface water, sediments and groundwater. This report does
not consider protective actions or emergency countermeasures to reduce
uptake in humans and domestic animals through the food pathway, for example
by feeding Prussian blue to cattle [26]. Strictly speaking, these are notremediation measures, since they only disrupt certain exposure pathways; theymay, however, be needed in intervention situations to supplement remedial
measures. Guidelines on their application have been published elsewhere [27].
This report is intended for individuals interested in the design, selection,
review or approval of projects to remediate sites containing dispersedradioactive contamination and provides a basic overview of the state ofknowledge for decision makers in government and at the community level andfor consultants.

1.3. STRUCTURE
This report is divided into seven sections. Section 2 outlines the nature ofdispersed radioactive contamination. Section 3 explains the fundamental
technical choices in making decisions about remediation and outlines basic
evaluation criteria influencing technology selection. Section 4 discusses themonitored natural attenuation and other non-intervention options. Section 5
describes the technical options for containing dispersed contaminants, while
Section 6 is devoted to removal techniques. A summary evaluation of thetechnological choices is provided in Section 7.

2. DISPERSED RADIOACTIVE CONTAMINATION
A variety of activities and accidents may result in dispersed (non-point)sources of radioactive contamination. For the purpose of this report, dispersed6contamination refers to the occurrence of concentrations of radioactive
isotopes distributed over a wide area, where complete removal and disposal of
the source is not practicable. Such contamination may present a hazard tohumans and the environment. Examples of sources include inadequate
practices for the management and disposal of radioactive waste, accidental
radioactive releases to the environment, major nuclear accidents, nuclear
weapon tests and incidents involving radionuclides at nuclear installations orother user establishments (such as hospitals and industrial and researchfacilities) [28]. A practice is any human activity that potentially introduces
additional sources of exposure or exposure pathways or extends exposure to
additional people or modifies the network of exposure pathways from existingsources so as to increase the exposure or the likelihood of exposure of peopleor the number of people exposed. Past practices that were not adequately
controlled have led to significant radioactive contamination of areas in
Member States. Some old facilities that processed radioactive material whenradiation protection criteria were not as stringent as they are today are sourcesof radioactive contamination. Radioactive contamination can also result
inadvertently from human activities involving processes in which naturally
occurring radioactive material (NORM) can be concentrated, in areas notnormally controlled by regulatory bodies, to levels beyond the limits set forpractices [6]. Such activities include conventional ore mining and processing(such as copper ore mining, phosphogypsum production or mineral sands
processing).
Contamination may take various forms and have different impacts onhuman health and environmental media; for example, a deliberate or
accidental discharge of material produced during mining tends to produce
dispersion plumes [28–31]. Conversely, accidents involving nuclear power
plants have, in the past, led to isolated but widely dispersed hot spots of activity
distributed across otherwise uncontaminated areas (e.g. Dounreay andChernobyl). The two cases are distinct and pose very different problems forworkers engaged in remediation.
International Commission on Radiological Protection Publication 60
stipulates that “remediation measures shall be justified by means of a decisionaiding process requiring a positive balance of all relevant attributes relating to
the contamination. In addition to the avertable annual doses, both individualand collective, other relevant attributes shall be assessed” [32]. Thus “remediation shall (a) reduce the doses to individuals or groups of individuals being
exposed; (b) avert doses to individuals or groups of individuals that are likely toarise in the future; (c) prevent and reduce environmental impacts from theradionuclides present in the contaminated area” [12]. A generic reference level
for aiding decisions on remediation is an individual existing annual effective

7

dose of 10 mSv from all sources, including natural background radiation. At
existing annual effective doses of less than 10 mSv, remediation would notnormally be justified on radiation protection grounds. An existing annual
equivalent dose of 100 mSv (inclusive of all existing contributions, including
doses due to the natural background radiation) to any organ will justify intervention under almost any circumstances [22]. Decisions on remediation willusually have to take into account other factors besides radiation protection
and, in particular, the views of those groups of the public that may be affectedby the situation and by any remediation action. In this context, the selection of
non-intrusive and less environmentally disruptive techniques will be preferable[17].
Finally, whether the contamination remains dispersed over time will
depend on the chemical and mineralogical characteristics of the contaminant,environmental transport processes and receptors. Physical, chemical andbiological processes leading to the concentration and fractionation ofradioactive species are discussed further in Section 4.

3. FACTORS INFLUENCING TECHNOLOGY SELECTION
3.1. FUNDAMENTAL TECHNICAL CHOICESThe objective of any technique used in a remediation project is either toremove or reduce the source term or to block the exposure pathways. This canbe achieved in a variety of ways and needs to be tailored to the contaminants
and pathways of interest. It may be necessary to use a suite of techniques toachieve the remediation objectives, especially for source term isolation or
removal.
In the case of dispersed contamination, a rigorous assessment of theactual and potential pathways is required to determine the optimal action. This
assessment begins with the identification and consultation of records, if
available. The historical assessment needs to be confirmed by a physical sitecharacterization, for example by walk-over gamma ray measurements.
Detailed sampling and analysis may be needed to more clearly identify hot
spots and to delineate materials that do not require further attention. In recent
years, a variety of strategies and techniques for efficient site characterization
have been developed [8, 33, 34].

8

Discrimination between zones containing different levels of activities is amajor challenge in reducing the areas and volumes of material to be considered[8, 35]. Once a site has been characterized, the decision makers are faced with
their first fundamental choice for the intended remedial action. They must
decide whether they will:
(a)

(b)

(c)

Leave the site undisturbed, while establishing a monitoring scheme fordetermining the evolution of the site. This option relies on naturalprocesses to prevent significant exposure. The entire process needs to becarefully monitored so that alternative action can be initiated if required.
Contain or restrict the mobility of the radioactive contaminants: thisinvolves immobilizing the contaminants inside the area in which they
already exist, reducing the potential for further migration or entry into
active pathways of exposure.
Remove the radioactive contaminants from the site, using an appropriate
treatment scheme: this involves extracting, concentrating and then safely
disposing of the contaminants at another location.

The three generic options that represent the fundamental technicalchoices for remediation can be summarized as monitored non-intervention,
containment and removal. Each of these fundamental technical choices will
direct decision makers to follow substantially different paths with regard totheir subsequent choices, actions and potential results, making available significantly different technological options for application (Fig. 3). In addition, sincea variety of remediation techniques exist for containing, reducing or removing
contamination, the technologies illustrated in Fig. 3 are grouped by the primaryemphasis of the technology into containment and separation or extraction. The
groupings are not necessarily mutually exclusive; for example, a barrier system
may be used to contain and extract a contaminant, and, in some cases, the useof a particular technique may occur on or off the site.
This choice, as discussed in Refs [3, 17], cannot be made solely on thebasis of scientific or engineering considerations. In addition to technicalconstraints, there may be a wide range of regulatory and socioeconomic
constraints on the selection of an appropriate remediation or disposal strategy[17, 36]. Regulations in Member States may favour certain techniques andprohibit or discourage others. International agreements may also preclude or
restrict some strategies. As has been discussed in detail elsewhere [29], the local
population may want to participate in the remediation decision making
process; public acceptability can be a major factor in selecting a particular
remediation technique. Active inclusion of the public will increase theirknowledge and awareness of the problem, increase acceptance of the9Recycleand reuse
Clearance
Solidification(chemical, thermal)

Radionuclides
Separationorextraction

Radionuclides

Segregation

Treatment
Clean soil

Clean water

Excavation

Solidification
andstabilization

Disposal
Isolation
Groundwatertable

Containment
Groundwater
pumping

FIG. 3. Classification of remediation technologies by function (adapted from Ref. [34]).

remediation technique selected for deployment and increase acceptance ofrestrictions on land use that may result. Participation may also enhance thepublic’s willingness to support the long term maintenance of remediation
measures and related installations.
Once these points have been clarified, measures may be chosen taking
the considerations described in Sections 3.2–3.8 into account.

3.2. TECHNOLOGY SOURCES
A wide variety of remediation techniques are now commercially available
or at the demonstration stage. Although most of these techniques are generic innature, others use proprietary formulations of reactants and other agents orapplications that are protected by patents or similar means. Since the field is
continuously developing, formal methods to assess the applicability and effectiveness of techniques have been developed [37]. Decision aiding tools makeclearer the evaluation criteria used to support the selection of a technique tomeet the remediation objectives [3, 5, 12, 38].
Approaches to selecting techniques vary from country to country. Somecountries regularly undertake technology assessments to help ensure that10

proposed projects are effective and efficient [39–49]. The findings typically are
made accessible in technology directories [50–54] or bibliographies [55, 56].
There are also international, semi-governmental, industry or researchcommunity sponsored initiatives [57–65]. Technology and technology suppliers
directories are available [66, 67]. Other Member States and organizations rely
on informal approaches, for example based on personal judgement by experts
and managers, to select techniques.
Once measurable remediation objectives have been established, several
factors affect the decision making process [3]. These basic evaluation criteria
include engineering and non-engineering considerations:
(a)
(b)
(c)
(d)
(e)
(f)

Effectiveness in remediating the contamination;
The costs associated with the remediation programme;
Occupational safety and health risks associated with the technique;
Prior experience with application of the technique;
Sustainability of any institutional control required;
Socioeconomic considerations.

3.3. EFFECTIVENESS
The use of the term ‘effectiveness’ in this section refers to the ability ofthe technique to contain, reduce or remove the contaminant in order to prevent
exposure of humans or undue harm to other properties of the site. Typically, an
overarching objective is that the remediation improves the radiologicalsituation by removing or reducing the source term or eliminating exposure
pathways that entail a health risk and that it is not detrimental to theenvironment [68]; for example, the annual dose is decreased, or the functionality of soils is retained or future land uses are not unnecessarily restricted.
Site specific considerations exert a significant influence on the effectiveness and efficiency of the chosen remediation method. Since the mineralogical and geochemical characteristics of the contaminant vary amongcontaminated sites, remediation methods are not universally effective andefficient. Methods to model and predict the effectiveness of techniques underconsideration have been developed [69, 70]. The anticipated performance of agiven technique can be simulated and compared with similar results from othertechniques to facilitate the selection [71]. The remediation action will becomplemented by a post-remediation assessment and monitoring programme
to ensure its efficacy [72] and that also may be part of any institutional control
required for residual contamination [13, 33, 73].

11

Steps have already been undertaken to incorporate remediation activities
into the ISO 9000/14000 quality management systems [71, 74]. Record keeping
is an integral part of quality assurance and/or quality control. It is essential thatrecords be kept of remedial actions undertaken, so that at any later point intime their performance can be evaluated against the original design. Having
comprehensive documentation available also facilitates intervention in theevent of unsatisfactory performance.

3.4. COST
The term ‘cost’ in this section covers the direct expenditure of funds
associated with the remediation technique. These include funds for licensing
fees for the technique, for equipment, labour and materials to deploy thetechnique, for design, construction management and treatability studies, for
operation and maintenance, and for monitoring and the disposal of residual
waste. Standard engineering cost principles can be applied to develop costestimates for remediation techniques. Major cost elements can be long termmonitoring, surveillance and maintenance. Depending on the time for which
institutional control is required, provisions need to be made for funding these
activities over periods of decades or even centuries.
Cost data for a wide variety of remediation techniques are available fromvarious sources; for example, the appendix of a recent IAEA document [17]provides an overview of remediation costs, drawing on national directories
such as the Historical Cost Assessment System (HCAS) [75] in the UnitedStates of America, which provides useful material for a relative cost assessmentof the techniques listed.

3.5. OCCUPATIONAL SAFETY AND HEALTH
The term ‘occupational safety and health’ in this section covers thepotential hazards and risks to workers involved in implementing theremediation technique [3]. Safety risks may result from accidents during
deployment. Health risks may result from workers being exposed toradionuclides. Occupational risks vary substantially among techniques and may
be an important consideration in selecting a given technique.
Information on safety and health risks can be obtained from standard
reference sources, regulatory standards, medical surveillance, safety studies,toxicological data and epidemiological studies [76, 77].

12

3.6. SECONDARY ENVIRONMENTAL IMPACTS
The implementation of a remediation project may result in a variety ofenvironmental impacts in addition to those resulting from the contamination
itself. Environmental risk involves adverse impacts on ecological receptors
located on or off the site due to a significant disturbance to the site ecosystem
and its surroundings as a result of remediation; for example, certain techniques,
such as the removal of topsoil or soil washing, may remove surface contamination at the cost of destroying the soil ecosystem.
Depending on the size of the site, an area larger than the actual contamination may be required for installations, the intermediate storage of waste andso on. The removal, transport and disposal of residual waste may result inenvironmental impacts and risks at locations other than those of the originalcontamination. There is, for example, little benefit in removing a contaminantthat is well fixed on a low volume of soil, only to produce a high volume of anaqueous waste with the contaminant in a soluble or mobile form. In addition,
the remediation techniques chosen may generate large quantities of secondary
waste and may pose risks of exposure of the public or operators that exceed therisks of quiescent contamination [78].
Environmental risk resulting from the implementation of remedialactions may also extend to possible impacts on natural resources such as
surface water, groundwater, air, geological resources or biological resources.
The potential for environmental risk may be an important factor in decision
making, since remediation technologies are more likely than others to produce
adverse impacts on ecological receptors, including habitat disruption, or to
generate damage to natural resources.

3.7. PRIOR EXPERIENCE
The term ‘prior experience’ in this section covers experience of implementing the remediation technique at other sites. It can be very useful to knowwhether the technique has been used successfully in the past. Information on
previous deployments is available from a number of sources, including vendors,
regulatory authorities, professional organizations, Internet databases, tradeassociations and publications (see Section 3.2).

13

3.8. SOCIOECONOMIC CONSIDERATIONS
The term ‘socioeconomic considerations’ in this section covers political,
social and economic factors that may influence the selection of a remediation
technique and its application at a site with dispersed radioactive contamination. The legal and institutional framework, prevailing socioeconomic
boundary conditions and public perceptions can influence the choice anddeployment of techniques to remediate sites with dispersed radioactive
contamination [17]. The level of public reassurance generally increases with thedegree of intervention and hence cost of the operation [29].

4. NON-INTERVENTION
4.1. BASIC CONSIDERATIONS
A decision not to intervene in site cleanup implies reliance on thecapacity of natural media (rocks, soils, sediments and groundwater) to retard
contaminant migration (i.e. natural attenuation) or on physical, chemical andbiological processes to reduce activity levels to below those of concern (i.e.
dilute and disperse). In either case, environmental monitoring is required toverify that such an approach is effective for the system under investigation. Itshould be noted that, ultimately, all remediation options that do not entail
complete removal of the contaminant source would de facto revert to thissolution if the half-lives of the radioactive species exceed a few hundred years.
It is also important to draw a distinction between those radioelementsthat occur naturally and those that do not, such as technetium and the transuranics. In the case of the former, reference may be made to the known
geochemistry of the element in a given environmental medium. This provides adegree of confidence in predicting future migration behaviour. For artificial
radioelements, experience is generally limited to laboratory data or small scale
field trials. The use of ‘chemical analogies’ in this context may be necessary but
is far from ideal [79].

144.2. DILUTION AND DISPERSIONRadioactive materials are discharged routinely into the air and surface
waters, both fresh and marine, from nuclear facilities worldwide. Prior to 1985,
dilute and disperse was also used for solid nuclear waste, although this practicehas since ceased following the London Convention [80] and OSPAR(Convention for the Protection of the Marine Environment of the North-East
Atlantic) Convention agreements [81]. Fewer restrictions apply to non-nuclear
radioactive waste, including effluents from mining and industrial activities
producing NORM. In the North Sea, for example, radioactive scale from pipe
work is removed in offshore facilities, milled and released with no further
treatment [82].
The effectiveness of dilution in aqueous media is critically dependent on
the speciation of the radioelement under the prevailing environmental
conditions [83, 84]. This will control factors such as solubility, adsorption tosurfaces, bioavailability and toxicity. Many radiologically important elements
may be concentrated by geochemical and/or biological processes, leading tosecondary sources of potential contamination [85–87]. Similarly, physicaldispersion of solids may not always be effective if the size and density of theparticles differ significantly from the ambient environment.
There is no doubt that, even where not proscribed by legislation, thedilute and disperse option is opposed by regulators, environmental groups andthe public at large.

4.3. NATURAL ATTENUATION
4.3.1.

Principles

The concept of natural attenuation has received a great deal of attention
in recent years [88–91]: it constitutes the least invasive approach to environmental restoration. The concept is not new; for example, it forms an integral
component of the design criteria for geological repositories that depend on
geochemical processes to retard radionuclide migration to the biosphere. It is
not entirely without financial cost. Reliance on natural attenuation requires
adequate monitoring, owing to the evolution of natural systems with time andour incomplete understanding of the processes operating. The effects of any
change in land use or in water abstraction would also need to be assessed,
hence the increased use of the term ‘monitored natural attenuation’ in theliterature [92–94].

15

A large number of processes can contribute to natural attenuation, as
discussed below. Figure 4 illustrates the effect of some of these processes on the
migration and concentration distribution of radionuclides. In order to beeffective, they must prevent or delay the arrival of a radionuclide at a receptor
until such time that it will have decayed to an insignificant level.
4.3.2.

Physical processes

4.3.2.1. Radioactive decay
The half-lives of radionuclides now present in the environment range
from microseconds to many millions of years. For higher members of thenatural series (234U, 235U, 238U, 232Th), together with some transuranics (e.g.
239Pu) and fission products (e.g. 99Tc, 129I), no substantial decay will haveoccurred even on the longest assessment timescale. However, many otherisotopes produced by nuclear fission (e.g. 60Co, 90Sr, 137Cs) or contained inindustrial NORM will not persist beyond a few hundred years. Clearly, it istherefore essential that a detailed radionuclide inventory be compiled before
deciding to adopt natural attenuation as a management policy at any given site.
The extreme fractionation between members of a decay series caused bychemical processing precludes the assumption of secular equilibrium in the
majority of cases [95].
4.3.2.2. Filtration
In most situations the dominant exposure pathway is via flowing water.
Resistate minerals (e.g. monazite, zircon and barite), other insoluble materials,
for example cement, or particulate matter on to which radionuclides havebecome bound may be retarded by filtration. This will depend on the relative
size of the particles and the pore distribution of the host medium, althougheven small colloids may be removed by fine grained clay matrices or fibrous
peat. In the case of aquifer transport, adequate characterization of the hydrogeological flow regime (permeability, hydraulic conductivity, heterogeneity,
fracture distribution) is a prerequisite for a quantitative assessment. Variably
saturated conditions and geotechnical issues also have to be taken into accountfor surface deposits.
4.3.2.3. Volatilization
Radon produced by decay of parent radium isotopes will escape fromwell ventilated soils or heaps, and hence the progeny will be subject to

16

C0

Concentration

Plug flow

Distance from source

C0

Concentration

Dilution by dispersion

Distance from source

Concentration

C0
Attenuation bydispersion and sorption

Distance from source

FIG. 4. Transport mechanisms effecting dilution and attenuation.

17

atmospheric dispersion. Methylated and permethylated forms of bismuth, lead,polonium and selenium (e.g. Ref. [96]) microbially generated in the subsurface
can also be volatilized.
4.3.3.

Chemical processes

4.3.3.1. Precipitation
Relatively few natural series radioelements and no artificial isotopes will
exist in sufficient mass concentrations to precipitate as a pure phase fromsurface, pore or groundwaters: the exceptions are uranium, lead and thorium.
Uranium is a relatively mobile element in the near surface zone, owing to
the stability of U(VI) aqueous complexes. However, it may be precipitated by
reduction to U(IV) or in the form of uranyl minerals, principally phosphates,
silicates, arsenates, vanadates and oxyhydroxides, several of which may occur
simultaneously at the same locality [97]. Around 150 such phases have been
identified and their formation represents the dominant fixation mechanism foruranium in weathering environments [97]. It follows that the amount of
uranium released to groundwaters or surface waters from these secondary
sources will depend on the solubility and dissolution rate of the phases as afunction of pH and water composition. Unfortunately, too few quantitative
data exist at present to allow predictive modelling [98, 99], an issue that needs
to be addressed.
Lead, for example, may precipitate as the insoluble sulphide galena (PbS)
that will incorporate 210Pb by isotopic substitution. Thorium occurs only in thetetravalent state and is substantially insoluble except at very low pH. Where
mobilized, for example in acid mine drainage (AMD), fixation occurs rapidly,
often within a few micrometres, via the formation of silicates or, in the absenceof silica, oxyhydroxides [100].
4.3.3.2. Co-precipitation
Radionuclides present at very low mass concentrations can neverthelessform solid phases by co-precipitation in mineral lattices. An important example
from both a nuclear and a NORM perspective is the high selectivity shown by
radium for barite, a mineral that has been very well characterized [101] and isalso exploited in a remediation context (e.g. Refs [102, 103]). It is likely thattransuranic isotopes would be similarly incorporated in uranium andlanthanide bearing minerals. There is evidence that this would be the case fromextensive laboratory studies [104], although it remains to be confirmed in thefield. Establishing the geochemical controls on migration of artificial radioele-

18ments is the major challenge to workers involved in the remediation of legacy
nuclear sites.
Co-precipitation on ferric oxyhydroxide flocs is an extremely efficient
removal mechanism for a large number of radioelements in solution. As thecontaminants tend to be released upon crystallization to goethite, the process isoften classified under the more general heading of sorption.
4.3.3.3. Sorption
In its strictest sense, sorption refers to the non-specific and reversible
uptake of ionic species at charged surface sites. Used loosely, it has come toencompass aspects of co-precipitation, ion exchange and a number of ion
specific interactions that are more appropriately termed complexation. The
distinction is not made here other than in the case of co-precipitation,
described above, as the latter clearly extends beyond the surface, resulting in
the formation of a defined mineral phase [105].
Clay minerals typically show a strong affinity for radionuclides in the
cationic form. Geological media with a high clay mineral content are therefore
more likely to effect attenuation. Adsorption and ion exchange would be
expected to play an important role in retarding the migration of soluble
monovalent and divalent ions. Examples include the pronounced retention of
caesium on zeolites (e.g. clinoptilolite) and the substitution of strontium forinterlayer cations in smectites. Surface sorption is an important transient for
multivalent ions in the formation of new mineral phases.
4.3.3.4. Complexation by organics
A number of radionuclides exhibit significant migration potential in thepresence of aqueous, low molecular weight organic compounds [106]. Equally,
however, immobile organic matter in the form of peat [107–109] or organic richhorizons in soils and sediments may provide an excellent substrate [107, 110]
for radionuclide retention. These phenomena have been studied extensively inthe context of ‘natural analogue’ studies for the performance assessment ofradioactive waste repositories [86, 111–113]. Uranium approaching percentage
levels has been reported in peat from Canada and northern Europe, whereas
iodine, often considered to be a conservative tracer in such assessments, hasbeen shown to be fixed in organic rich lacustrine deposits [114].

19

4.3.4.

Biological processes

Biomineralization, ‘biosorption’ and microbially mediated phase transfer
also can effect attenuation. These processes are discussed in more detail in
Sections 5.5–5.7, 6.8 and 6.11. The complex biogeochemical processes effecting
the fixation or mobilization of metals, including radionuclides, in various typesof soil ecosystems have been studied with increased intensity in the aftermathof the Chernobyl accident [115] and in other remediation contexts [116].
4.3.5.

Definitions and applicability

Whether to intervene or to rely on natural attenuation can only be
determined on a site by site basis [117]. Factors militating against intervention
include:
(a)
(b)
(c)
(d)
(e)
(f)
(g)
(h)
(i)

The areal extent of the contamination;
The accessibility of the site;
The proximity to sensitive receptors;
The radionuclide inventory;
The time frame;
The presence or absence of co-contaminants;
The chemical and mineralogical characteristics of the material;
In the case of surface deposits, the geotechnical stability;
The transmissivity of the host medium.

A comprehensive site investigation programme is essential to determine
these factors.
The degree of confidence that can be ascribed to natural attenuation inpreventing harmful exposure or environmental damage is proportional to thelevel of characterization of that site. Developing an understanding of thephysical, chemical and biological processes operating is more crucial in the caseof natural attenuation than if the contamination were to be removed physicallyfrom the site.
A decision to apply monitored natural attenuation (Fig. 5) as thepreferred management strategy will invariably be made by considering a
combination of scientific, economic and political criteria. Ideally it should bebased on a prior risk analysis of the specific site and follow an establishedtechnical protocol [118]. Models exist for evaluating the likely performance ofsuch systems, for designing the monitoring programme and for the interpretation of monitoring data [112, 113]. Given a backdrop of scarce resources,
various initiatives are under way to promote the acceptance of natural

20

Groundwater flow

Critical group

Groundwater flow

Backgroundmonitoring
wells

Shallow monitoring Deep monitoring well
well

Criticalgroup

Contaminatedarea
Shallow aquifer
Water well
Deeper aquifer

FIG. 5. The principle of monitored natural attenuation.

attenuation as a part of a cost effective and environmentally sound solution forradioactively contaminated sites worldwide (e.g. Ref. [119]).
It may be possible to modify some property of the media hosting thecontamination in such a way that the retention capacity of the site is substantially enhanced. Examples include the addition of certain minerals, revegetation (see Section 5.6) and the creation of wetlands [120] (see Section 5.7).
Strictly, any attempt to modify the characteristics of a contaminated site
represents a form of intervention; for this reason, such measures are discussed
further in the next section.

4.4. ALTERNATIVE LAND USES
When extensive areas have been contaminated, many of the discussed
remediation methods may be too expensive to carry out or too intrusive. In
particular, when the land was used for agricultural purposes, alternative uses
may need to be considered. Such alternative uses may range from switching todifferent crops to turning to completely different uses, such as parkland.

21

When selecting an appropriate countermeasure for a given area andcontamination problem, the optimum solution will depend, apart from radiological criteria, as much on economic, social and political factors as on soundscientific considerations [121].
Many studies have been targeting possible agricultural countermeasuresin response to concentration levels in food and agricultural crops exceeding theapplicable standards. Most studies have been conducted to test the effect ofdifferent physical and chemical countermeasures. However, information on thelong term effect of countermeasures, and especially of a change to non-foodcrops, is still scarce.
When investigating alternative crops, the principal questions to beaddressed are:
(a)
(b)
(c)
(d)
(e)
(f)

Can an alternative crop be found that is suited to the climate and soil
conditions prevailing in the contaminated area?
What is the fate of the radionuclide in the cultivation system and along
conversion routes?
How does the radionuclide in question behave during biomass processing
and what is the expected radionuclide concentration in the end products?
What is the exposure during biomass cultivation and processing?
Would production and utilization of the alternative crop be economicallyfeasible?
What are the overall prospects for the chosen alternative crop as analternative land use for large contaminated areas?

In order to understand the fate of the various radionuclides and theirdistribution in products, residues and waste, one needs to know the various
radionuclide fluxes. These depend on the initial deposition levels, crop accumulation factors, which in turn depend on plant and soil characteristics, and theradionuclide accumulation in the produce (e.g. wood, rape or beetroot).
Whether residues and waste need to be treated as radioactive waste depends
on the radionuclide concentration and the applicable exemption limits.
Crops used for liquid biofuel (oils, alcohol) production, such as rape,
wheat, sugar beet, barley, potatoes and winter rye, may be suitable alternative
crops.
The data in Table 2 indicate that crops with a low transfer factor (TF) tothe useable product can be found and that the resulting liquid biofuels arealmost free from activity, and that 137Cs levels in the waste and residues are
generally of no concern.
Examples from Belarus, however, show that caesium levels in oil cakefrom rapeseed oil (~2000 t/ha) and the pulp and vines from sugar beet (~4000 t/ha)

22

TABLE 2. CAESIUM TRANSFER FACTORS TO DIFFERENT PLANTPARTS OF SOME POTENTIAL BIOFUEL CROPS [122]CropSpring wheat

Plant component

Caesium TF (10–3 m2/kg)

Straw

0.23–0.36

Grain

0.13–0.16

Straw

0.27–0.44

Grain

0.08–0.18

Straw

0.43–0.60

Grain

0.17–0.29

Green mass

0.33–0.81

Straw

0.38–0.92

Seeds

0.27–0.66

Brassicaceae

Seeds

0.037–3.4a

Peas

Seeds

0.69–1.25

Straw

0.82–1.45

Leguminosae

Seeds

94 (12–750)a

Sugar beet

Root

0.43b

Root crops

Root0.025–11 (1.1–110)a

Green vegetables

Leaves

0.07–4.86a

Leaves (peaty)

260 (25–2700)a

Winter wheat
Rye
Spring rape

Sunflowers

a
b

Straw

1.48–2.88

Seeds

0.43–0.82

From Ref. [123] (average and 95% confidence limits).
From Ref. [124].

may be too high for use as animal fodder and for incineration and that they may
have to be disposed of as radioactive waste [125]. This, however, applies
primarily to cases of high contamination levels, since only about 3% of the total
137
Cs and 6% of the total 90Sr taken up will be potentially involved in the soil–
plant–fodder–animal–human chain [126]. Moreover, oil cake and sugar beet
pulp usually constitute only 10% of fodder, and hence generally animal fodder
can be used without restriction [126]. The actual cost of liquid biofuel typicallyis significantly higher than that of fossil fuels, and hence a price subsidy or othermeasures may be needed [125, 126].

23

On the other hand, the production of rapeseed and processing to edible
rapeseed oil are profitable technologies [122, 126], and the levels of caesium
and strontium in the rapeseed oil after three filtrations and bleaching are below
the detection limit [126].
The feasibility of valorization of contaminated land by willow shortrotation coppice (SRC) for energy production has been addressed in various
studies [125, 127–129]. Coppicing is a method of vegetative forest regeneration
by cutting trees at the base of their trunk at regular time intervals. Fast growing
species of the Salix genus (willows) are frequently used in a coppice systembecause of the ease of their vegetative reproduction and the large biomass
produced. The harvested biomass is converted into heat or power (with anappropriate off-gas treatment). As such, this non-food industrial crop is apotential candidate for the valorization of contaminated land that has use
restrictions. SRC may be preferred over traditional forestry since revenues
come sooner after establishment and more regularly (every 3–5 years). SRC
yields are also high on good agricultural soils, and its use is not a drastic change
in land use; SRC is easy to introduce and it is easy to return the land to the
production of food crops. SRC may also be considered as complementary toforestry, given the different culture requirements of both vegetation systems.
Forests perform well on sandy soils, whereas SRC requires soils with asufficient water retention capacity. SRC has additional potential advantages ina contamination scenario: since it is a perennial crop, dispersion of radionuclides will be limited. Harvest can be in winter, when the soil may be covered bysnow, resulting in radiation protection of the workers. Finally, SRC cultivation
is not too labour intensive, which is also an advantage with respect to exposure.
Willow SRC may be a suitable rehabilitation tool for highly contaminatedland, but only if the radionuclide levels in the wood are below the exemption
limits for fuelwood, if the average yearly dose received during coppicecultivation and coppice wood conversion is acceptable, if SRC can be grown
successfully in the contaminated territories (soils, climate), if the cultivation of
SRC is technically feasible and if SRC production and conversion are economically profitable.
It has been shown [125, 127–129] that for soils with a medium to highfixation (finer textured soils) and sufficient potassium availability, the TF ratioof concentration in plant biomass to concentration in soil is <10–5 m2/kg, andwood can be safely burnt and the ashes can be disposed of without concern. For
light textured soils, however, with a low radiocaesium fixation and low soil
potassium, the TF to wood is around 10–3 m2/kg, and concentrations in wood
may be elevated enough that the prevailing exemption limits are reached.
Given that TFs for common forestry and for straw of winter wheat and rape arecomparable, the same applies for burning wood or straw for energy.

24SRC has generally a high annual yield of about 12 t/ha, but sandy soils are
only suitable for SRC production if well fertilized and irrigated. Only during theconversion phase and when burning highly contaminated wood (3000 Bq/kg) dodoses in the vicinity of ash collectors exceed the level of 1 mSv/a for a memberof the general public [125]. Contributions from other possible exposure
pathways are negligible (external exposure during cultivation and transport,
inhalation dose in the combustion plant and doses to the public following woodburning).
Crop yield and the capital cost of the conversion units are among themost important parameters affecting system profitability. At the production
site, a minimum yield of 6 t/ha/a is required for Belarus production conditions
and of 12 t/ha/a for western European conditions, if all other parameters are
optimal [125, 128]. Heating schemes may be a viable option for wood
conversion in Belarus, whereas electricity generation schemes are not.
Subsidies would be required in Europe to make wood conversion economicallyfeasible. It has also been concluded that the existence of a contamination
scenario does not necessarily hamper the economic viability of the energy
production schemes studied. The cost associated with the disposal of contaminated ashes was estimated as less than 1% of the biofuel cost and will not affecteconomic feasibility.
Forestry can also be considered to be an adequate alternative land use[130, 131]. Soil to wood TFs to coniferous and deciduous wood are around
10–3 m2/kg [132] and are hence comparable with the TFs to willow woodobserved for low fertile soils with limited caesium fixation. They are highcompared with the TFs observed for willow in finer textured soils and soils withan adequate potassium status. Moreover, the annual biomass increase is only 6t/ha for forests and may attain 12 t/ha for SRC grown on soils with an adequate
water reserve and fertility status. SRC may hence be a more promising land use
option on these types of soil than traditional forestry. On soils with a low water
reserve (e.g. sandy soil), however, willow yield without irrigation is too low to
be economically feasible, and forestry may hence be the preferred option [129].
Fibre crops are also potential alternative crops for agricultural land withrestricted use. Potentially suitable crops are the annual fibre crops hemp
(Cannabis sativa L.) and flax (Linum usitatissimum L.). Hemp and flax are wellknown arable crops that have been cultivated for centuries. Ukraine has alegacy of flax and fibre hemp cultivation, but in Belarus there is only some flax
production. Since the early 1990s the acreage for production of flax and hemp
has declined dramatically in Ukraine. Establishment of fibre crops on contaminated arable land is generally of no radiological concern [133]. The TFs
observed to hemp fibres are a factor of 4 to 50 higher than the TFs observed toflax. Cultivation is hence generally restricted to not too contaminated areas25

(<1000 kBq/m2). For both crops it holds that contamination levels in the waste
products (oil seed cake, chaff, ash after burning of straw) may, however, be highenough that they should be considered as radioactive waste. The economics ofthis land use has not, however, been investigated [133].
The introduction of alternative crops in a contamination scenario may be
a feasible and adequate remedial option. Although there are some scenarios inwhich energy production from SRC and potentially other alternative crops on
contaminated arable land is radiologically safe and economically feasible,
installing this cultivation system on a large scale requires extensive logistics,
infrastructure and initial investment. Implementation is likely only to be
successful with adequate political support.
There are a number of additional types of alternative land use, such as thecreation of parkland. Such measures, however, would largely be administrative
and would amount to ‘institutional control’ and therefore are not consideredfurther in this report.

4.5. AGRICULTURAL COUNTERMEASURES
Many studies have been concerned with possible agricultural countermeasures in response to concentration levels in foodstuffs and agricultural
crops exceeding the permissible levels in the wake of the Chernobyl accident;
for example, the efficiency of countermeasures for radioactively contaminatednatural and agricultural ecosystems was evaluated recently in the context ofChernobyl in Ref. [134]. To this end, a database of 5261 experiments carriedout during 1987–1999 and their respective results was compiled by participantsfrom Belarus, the Russian Federation and Ukraine [134]. The main evaluation
criterion was the efficiency of experimental treatments in reducing radionuclide concentrations in final products as compared with untreated controls. It is
important to note, however, that the majority of countermeasures do notintend to influence soil or groundwater concentrations, but aim to breakexposure pathways.
Countermeasures can be based on a selection of crops that exhibit smaller
radionuclide uptake than crops used previously, on food processing to reduceradionuclide contents or on choosing non-food crops, resulting in either case ina produce from the contaminated land that is radiologically acceptable [25, 121,135]. Impacts on the dose received by people and on the ecology and economyof the affected area may vary enormously: a change in crop variety will have amuch smaller impact than more radical changes such as the substitution ofvegetables by cereals or changing from arable land use to animal husbandry.

26

Assessments have shown that substituting crops and fertilization are themost effective countermeasures in plant production. The efficiency of countermeasures, expressed by the reduction factor of radionuclide concentration in
final products, was found to be of the order of 3 to 9, depending on the soil andindividual crops [134]. Substituting crops may not be expensive, but its viability
depends on a variety of economic conditions; for example, on contaminatedsandy soils in Belarus it would be possible to grow rye that has an acceptableconcentration of 137Cs, but market demand for rye is poor and recently has
been tending to zero. Another example are the impediments for the utilizationof sugar beet: there is no sugar processing industry in the contaminated area in
Belarus and no agricultural infrastructure, such as suitable farm machinery, forcultivating the plants.
Fertilizer application will suppress the uptake of certain radionuclides,
mainly due to competitive effects. Thus potassium dosages will generally
decrease the soil to plant transfer of 137Cs, certainly when the soil is low in
potassium. Reported reduction factors have varied between studies, but overall
reduction factors ranging between 1.1 and 5.0 have been obtained [25, 134,
136–138]. The efficiency of potassium additions strongly depends on theexchangeable potassium content in the soil. For soils with a low to optimal
potassium content, high dosages of potassium fertilizer are very effective andprofitable. For soils with a high potassium content, only moderate dosages ofpotassium fertilizer are recommended to replace the potassium removed withcrop yields [139, 140].
The behaviour of 90Sr and its uptake by plants are controlled by its
similarity to calcium. Many investigators have found a significant correlation
between strontium TFs and the reciprocal of the exchangeable calcium content[141]. Consequently, much of the research and actions to reduce strontium
uptake by plants has centred on the use of lime as a soil based countermeasure.
The use of lime has reduced strontium uptake by up to 40%, the use of limed
compost by up to 60% [141]. Generally, the reduction factor of radionuclide
uptake by agricultural crops varies widely, from 1.1 to 3, depending strongly on
the initial soil pH. The liming effect is most pronounced for acid soils [27].
Countermeasures that aim to provide the optimum (from the plant
production point of view) rates of fertilizer application are the most viable,
since the investment on fertilizer is paid back in the form of additional crop
yields, and frequently profits are made. It has to be noted that the addition of
nitrogen fertilizer should be moderate, as high dosages appear to stimulate theaccumulation of 137Cs and 90Sr in plants [142–145]. Phosphorus dosages should
be in accordance with crop responses and the phosphorus content of thetreated soil, and should be crop specific.

27

The humus content of a soil is important because of its tendency to form
co-ordinate bonds with calcium and strontium, which are stronger than thebinding by ion exchange sites on soil minerals. Organic matter addition has
resulted in strontium TF reductions by a factor of 1.2 to 7. The latter value wasobtained after the addition of 15% organic matter to a sandy soil [141]. Field
experiments on a podzoluvisol (loamy sand) soil in Belarus that increased
humus content from 1.5% to 3.5% resulted in a reduction of 137Cs and 90Sr
activity in perennial grass by a factor of 2 [142, 144].
Chemical amendments, such as zeolites, ammonium hexacyanoferrate(AFCF) or clay minerals, also reduce radionuclide uptake by plants, since theradionuclides are trapped and so rendered less available for plants [146–149].
A reduction factor of 4.6 in 90Sr transfer has been obtained for a sandy soil after
the addition of 1% zeolites [149], and a factor of 25 by applying 10 g AFCF persquare metre [146]. However, the investigation of zeolites and clay
amendments in field trials on a loamy sand soil in Belarus resulted in only arather low reduction of activity in cereals [150]. Taking into account the highcost of transport and application of amendments at reasonable rates, forexample 15 t/ha, this type of countermeasure is generally not affordable inconventional agriculture.
A more radical improvement of private hay land and meadows in all
Chernobyl affected rural settlements of Belarus, the Russian Federation andUkraine (where the total annual dose is above 1 mSv or the activity per unit
area exceeds 555 kBq/m2) is recommended. This countermeasure combines theliming of acid soils, fertilization (including the basic application of organic fertilizers), destruction of old turf, sowing of new grass stand and regulation of soil
water (drainage), if needed; for example, radical meadow improvement has
resulted in a reduction of grass activity by a factor of 1.7 to 3.5 [151], but otherapplications have achieved reduction factors for 137Cs of up to 16–20 [134]. The
reduction factor of surface meadow improvement is lower, and is 3.5 on
average.
Although strictly speaking not remediation techniques, certain livestock
management measures are effective in reducing public exposures. Suchmeasures include feeding complexants, such as Prussian blue, to dairy animals
to prevent 137Cs transfer into the milk, or changes in pasture or fodder atcritical times to reduce uptake. Achieved reduction factors vary widely
between 2 and 15 [134].
Similarly, use and access restrictions for forests are more emergency
responses than remediation measures.
Food processing can significantly reduce radionuclide concentrations in
products. The efficiency depends on the type of processing and varies widely,

28

TABLE 3. EFFECTS OF VARIOUS COUNTERMEASURES ON RAPE
RADIOACTIVITY [126]Activity reduction in rape products
137

90

14%

42%

Application of N90P90K180 fertilizer

42%

27%

Liming to 6 t/ha + N90P90K180 fertilizer

45%

59%

Variety selection

2.5 times

3.0 times

Rapeseed oil processing (crude oil)

250 times

600 times

Cs

Liming to 6 t/haSrremoving 50–98% of the 137Cs or 90Sr during the production of butter or casein
from milk [134].
The relative efficiency of different agricultural countermeasures can beseen from the experiments in which rape was grown on radioactively contaminated land in Belarus (Table 3). The effect of liming is mainly due to a rise of
the soil pH and hence the increased availability of exchangeable calcium.
Choosing a rape variety with less uptake offers activity reductions of up to
three times. The most efficient removal of activity is offered by oil processing,resulting in a reduction of up to 600 times. Concentrations of radionuclides
after a three stage filtration and bleaching are below the limits of detection.
The combination of oil seed processing with several agricultural countermeasures therefore allows the production of food grade oil practically freefrom radionuclides and produces a valuable protein by-product (cake as animal
fodder) with permissible concentrations of radionuclides [126].

5. CONTAINMENT
5.1. ENHANCED ATTENUATION
Although contaminated media sometimes provide sufficient attenuation
capacity, normally these attenuation mechanisms must be enhanced throughtechnical measures. Enhancement of attenuation may be targeted at particular
exposure pathways; for example, plant uptake may be minimized or blocked toprevent contaminants from entering the food chain [27, 152].

29

Simple ploughing or deep soil mixing is not an efficient means of reducing
direct surface gamma exposure [153], as such an approach will result in adispersal of radionuclides over a larger area, thereby increasing the volume ofcontaminated soil.
Studies subsequent to the Chernobyl accident found that deep ploughingwith digging, combined with liming and potassium fertilizer application, candecrease caesium and strontium transfer from soil to plants by a factor of 3 to 4
[23, 150]. The objective of deep ploughing is to skim off the upper 0–5 cm
contaminated soil layer and burrow it beneath the turned over arable layer(30–50 cm), thereby preserving most of the soil fertility. Subsequent cultivation
practices have to be limited to shallower depths to prevent the contaminatedsoil layer from being dug up or roots from reaching this layer. This quite costeffective countermeasure had only limited application after the Chernobyl
accident because of the thin humus horizon of the predominantly light textured
soils in the region [143, 150].
Changing the pH and redox conditions in contaminated zones canenhance attenuation [119, 120, 154–157], particularly in situations in which
treatment is otherwise difficult, such as the presence of fractured rock [152].
Oxidation or reduction can be achieved by injecting aqueous solutions ofappropriate agents, or by bubbling gases through the contaminated zone [156,158]. Long term sustainability is uncertain, and competing geochemical
processes need to be evaluated carefully [152, 159]. Environments withrelatively low redox potential and high organic matter content (e.g. wetlands)
tend to trap metals naturally [120], a property that can be utilized (Section 5.7).
The number of sorption sites may be increased by adding clay or zeolites[160] to soils. The addition of reactive minerals, such as lime, apatite [160–162]
and its derivatives, such as bone meal [163, 164], may lead to immobilization
through the formation of sparingly soluble mineral phases incorporating thecontaminating radionuclides.

5.2.

LOW PERMEABILITY BARRIERS

Low permeability barriers, which include slurry walls, sealable-joint sheet
piles and curtains created by injection grouting [165–167], interrupt water
flowpaths and may delay the migration of radionuclides. The construction ofsoil–bentonite slurry walls and cement–bentonite slurry walls to serve as low
permeability barriers is a well established technique, has been demonstrated incivil construction and has been transferred to provide containment at contaminated sites [168]. The efficacy of low permeability barriers, however, depends
critically on strict quality control during construction. Such low permeability

30

Slurry wall

Contaminant plume

Reactive wall

Sheet pile
wall

FIG. 6. Sketch of a permeable reactive wall in combination with a Funnel-and-Gate
system.

barriers may be used in isolation or as a component of a suite of techniques(Fig. 6).

5.3.

PERMEABLE REACTIVE BARRIERSThe use of permeable reactive barriers or walls is distinguished fromoutright containment by the fact that the contaminant carrier as such (i.e. thegroundwater) is not prevented from movement [169–172]. The objective israther to remove the contaminants from the mobile phase. Permeable reactive
barriers are installed by excavating a portion of the aquifer, disposing of the
excavated material and replacing it with a permeable material designed to reactwith the contaminant and remove it from the flowing water (Fig. 6). The
advantages over pump and treat systems are that no active pumping or process
operation and maintenance is required, thus reducing energy and operationand maintenance costs, no treatment sludges are produced, thus reducing waste
disposal costs, and no surface facility is required, which allows the land to be
returned to productive use. The systems typically rely on the natural gradient of
the groundwater table as the driving force. The barrier material must be
designed to remain reactive for periods of many years to decades. Furthermore,31the barrier permeability must be sustained throughout the duration of thegroundwater treatment. The performance of permeable reactive barriersystems must therefore be monitored so that corrective action can be taken
when required.
Permeable reactive barriers have been designed and implemented for thetreatment of dissolved metals [169, 173, 174], acid mine drainage [175–177],
radionuclides [178–181] and dissolved nutrients [182–184]. Contaminantremoval can be effected in a variety of ways [185–187]. Treatment processes
include adsorption [188, 189], simple precipitation [190], adsorptive precipitation [183], reductive precipitation [169, 176, 179] and biologically mediatedtransformations [175, 176, 182].
Changing the redox state can be a very effective method of immobilizing
certain radionuclides (e.g. uranium and technetium). These radionuclides have
two or more oxidation states, and the more reduced oxidation states are lessmobile; for example, reduction of the hexavalent uranyl ion UO22+ to thetetravalent U(IV) state results in the precipitation of sparingly soluble precipitates, including UO2(s) or mixed U(VI)–U(IV). Zero valent iron is an abundantand inexpensive reducing agent that has been observed to reduce andprecipitate uranium and technetium in laboratory studies [178, 179, 191–193].
The oxidation products generated (e.g. ferric hydroxides) can provide a highcapacity sorption substrate also for non-redox sensitive species [194], but theirlong term stability in relation to changes in redox conditions has to be carefully
evaluated [195].
Permeable reactive barrier systems containing zero valent iron have been
installed for the treatment of uranium, technetium and other metals [179, 180,
196, 197]; these barriers demonstrate excellent removal of uranium andtechnetium. Examination of the reaction products has been conducted at aseries of sites of permeable reactive barriers [179, 198, 199]. Although theresults of these characterization studies are inconsistent, all the reports indicatethat a portion of the uranium entering the barrier system is reduced to U(IV),whereas some portion may remain in the U(VI) oxidation state. Other metals
commonly associated with uranium mine waste, including arsenic,
molybdenum, selenium, vanadium and zinc, are also removed from the groundwater, possibly as reduced phases (e.g. V2O3) or as sulphide minerals (As2S3,
ZnS) [180, 200].
Organic reductants, such as sawdust, have also been used to promote thereduction and precipitation of uranium. Passive treatment systems containing
organic carbon have been used to remove both uranium and nitrate fromgroundwater at sites where these two constituents coexist as a result of releases
from nuclear weapon production facilities [201].

32

Sorption can remove contaminants from groundwater and can maintainlow concentrations of radionuclides. Sorptive materials that have been
evaluated or deployed in permeable reactive barrier systems for treating radionuclides include zeolites (e.g. clinoptilolite [202]), phosphate based adsorbents(e.g. bone char apatite [196] and Apatite II [203]) and hydrous ferric oxides(e.g. amorphous ferric oxyhydroxide (AFO)) [188, 196].
The majority of the reactive barriers installed to date have been
continuous barriers installed across the entire width of the plume. Contaminant
fluxes also can be focused on the reactive barrier by an array of non-reactive
barriers, such as slit or slurry walls [166], to form a Funnel-and-Gate system[204, 205]. Funnel-and-Gate systems reduce the physical length of thetreatment portion of the barrier and prevent contaminants from circumflowing
the treatment zone. The volume of reactive material required to treat contaminated groundwater is determined by the contaminant concentrations,
groundwater geochemistry and flow rate. For many contaminant plumes, thevolume of reactive material will be similar, whether a continuous barrier or
Funnel-and-Gate configuration is employed. Since the installation ofcontinuous barriers is typically less expensive than that of Funnel-and-Gate
systems, this installation technique has been preferred. Furthermore, because
Funnel-and-Gate installations focus the flow to across a small cross-sectionalarea, there is greater potential for clogging by the formation of secondary
precipitates.
Depending on the reactive material to be used, deployment techniques
may include injection wells (for grouts, gels and soluble reactants) or trenches(see Fig. 7) cut by a suitable excavator (for grouts and particulate material suchas granular iron, sawdust, etc.). Development work on efficient methods toemplace reactive barriers with minimal disturbance, even in awkward places, isongoing. Adaptation of more novel civil engineering techniques, such as
directional or horizontal drilling, the use of guar gum slurries for barrier installation [177], hydraulic fracturing [206] and jet grouting techniques [207], can beused for the emplacement of barriers at depths beyond the capabilities of
conventional excavation techniques (Fig. 8).
In December 1998 a reactive barrier incorporating a novel wall andcurtain design was installed at the Atomic Energy of Canada Limited (AECL)
research facility in Chalk River, Ontario, Canada, about 200 km west of Ottawa[202]. The wall and curtain design directs contaminated groundwater to thetreatment media, and excludes the overlying uncontaminated groundwaterfrom the treatment zone (Fig. 9). The treatment zone in the wall and curtain
barrier contains clinoptilolite, a naturally occurring zeolite mineral thatremoves 90Sr from the plume via a sorption reaction. Preliminary monitoring of

33

(a)

(b)

FIG. 7. (a) Continuous trenching machine used to install the 46 m long, 7.3 m deep and
0.6 m wide granular iron permeable reactive barrier; (b) simultaneous excavation andreplacement of aquifer material with granular iron as the horizontal trencher advances[193].

34

Rig fordirectional drilling

Subsurface
contamination

Injected grout,etc.

Groundwater

FIG. 8. The principle of directional drilling and grouting.

the barrier performance has suggested excellent removal of 90Sr, from initialconcentrations of 0.1–100 Bq/L to below the Canadian drinking water limits[189]. The ongoing monitoring programme has shown that the barrier has
treated more than 1.85 × 107 Bq of 90Sr.
A three component Funnel-and-Gate reactive barrier system wasinstalled at a former uranium upgrading site at Fry Canyon, Utah, USA, in

Curtain of reactive media
Sheet pile cut-off wall
Level control manhole
CaptureZone

To creek

FIG. 9. Trench and gate system [202].

35

1997 [196]. One of the permeable reactive barrier components contains AFO,
and a second contains bone char apatite. Both of these components areintended to assess the potential for uranium treatment by sorptive processes.
The third component contains zero valent iron, intended to reduce andprecipitate uranium as a reduced uranium oxide. The performance of thebarriers has been monitored over three years and 1650 pore volumes. The zero
valent iron component demonstrated the greatest removal efficiency (C/C0 =0.007), followed by the AFO system (C/C0 = 0.271) and the bone char system(C/C0 = 0.412).
Computer simulations conducted using reactive solute transport models
can be used to determine design parameters for barrier installation, to predictthe potential for barrier clogging and to assess the potential benefits of barrier
performance. The performance of a reactive barrier installed at the Elizabeth
City US Coast Guard Support Centre [208] was simulated using the reactive
solute transport model MIN3P [209]. Comparison of the simulation results withsubsequent measurements showed good agreement (Fig. 10). The performanceof the permeable reactive barrier installed at Monticello Canyon, Utah, USA,was simulated using the PHREEQC model [180].
The limitations on permeable reactive barrier performance and lifespaninclude constraints on the reactive material longevity and the barrier permeability. Of these concerns, the potential for barrier clogging and the permeable
reactive barrier evolving into an impermeable reactive barrier is the mostsignificant. Since the total mass of contaminant that accumulates in the barrier
is modest, the principal precipitates resulting in clogging are the products ofreactions between the barrier material and the major ions present in the water,or between the barrier material and the water itself. The use of zero valent iron(Fe0), the most commonly used reactive material, results in the reduction ofwater and an increase in the pH to between pH10 and pH11. This increase in
pH favours the precipitation of carbonate minerals, principally calcite (CaCO3)
and siderite (FeCO3). Over periods of several years to decades, the accumulation of these precipitates potentially may be sufficient to reduce the porespace of the reactive material and limit barrier permeability. Reactive barrier
technology has evolved recently, and the oldest barriers are now approaching
ten years of operation. Clogging to a degree that is sufficient to impair barrier
performance has yet to be observed, although long term monitoring
programmes are required to assess this concern.
The long term fate of the reactive barrier after remediation is complete orafter the barrier becomes ineffective depends on the nature of the contaminant
and on the characteristics of the barrier. Concerns include the potential forremobilization of contaminants retained in the barrier and the potential for
clogging in the barrier to alter natural groundwater flow conditions. In many

36

(a)

Barrier4

Depth (m)

pH

11

5

7.5

6

6

Flow
7
0

1

2
Distance (m)

3

4

(b)

Volume fraction

0.02
CalciteCa–Mg carbonate
Siderite
Fe(OH)2(s)Ferrihydrite
Mackinawite

0.015
0.01
0.005
0
1.5

1.75

2.0
Distance (m)

2.25

2.5

FIG. 10. Simulated heterogeneous reactions at a permeable reactive barrier at the
Elizabeth City US Coast Guard Support Centre [209].

barrier systems, the contaminant is converted to a form that is stable in thegeochemical environment that prevails in the aquifer. Furthermore, because
the mass of contaminant is small relative to the mass of the barrier material, theresidual barrier material may be classified as non-hazardous. In these systems,
it may be acceptable for the barrier to remain in place. In other cases, the massof contaminant may exceed soil guidelines, the contaminant may have thepotential for remobilization or the contaminant may be sufficiently hazardous
to warrant excavation of the reactive material and placement in a secure waste
disposal facility. In these cases, excavation of the barrier, or a portion of thebarrier, may be required.

37

Although considerable research on the performance of reactive walls is
continuing worldwide, some techniques have reached commercial maturity[210, 211].

5.4. IMMOBILIZATION
Immobilization, unlike physical containment, is intended to affect thecontaminated material itself. The objective of immobilization is to change thecontaminant form into one that is less susceptible to migration. Two basic
options can be distinguished: in situ and ex situ treatment [212, 213].
In situ treatment effects immobilization without the contaminatedmaterial being removed. Three major methods to effect in situ immobilization
can be distinguished, based on chemical, biochemical or thermal treatments.
Chemical immobilization is based on the injection of a variety of grouts[214] or on changing pH and/or redox conditions in the groundwater, forexample [120, 154, 155, 215]. These grouts can be based inter alia on ordinary
Portland cement (OPC), water glass (sodium silicate), gypsum or organic
polymers, for example acrylic or epoxy resins. The suitability of immobilizing
agents via injection depends largely on the hydraulic properties of the contaminated material. OPC and epoxy resins typically have a high viscosity, whilewater glass and gypsum solutions, or acrylic acid suspensions, can be made up
with viscosities equal to that of water. The long term stability of the polymer
stabilized material has to be carefully assessed. Breakdown products
containing functional groups, such as carboxylic or phenolic groups, may
actually act as a vehicle to facilitate transport of radionuclides.
Injection of chemical reductants, including calcium polysulphide, hasbeen used to promote contaminant reduction and precipitation within aquifers.
Contaminants that are well suited to remediation using this approach include
metals with a lower solubility under reduced conditions. Injection techniques
have been used to treat Cr(VI), through reduction to Cr(III) and precipitationof Cr(III) hydroxides. In situ redox manipulation (ISRM) [155, 215] is avariation on a chemical injection system (Fig. 11). When using ISRM, a strong
reductant is pumped into the aquifer, converting oxidized Fe(III) bearing
minerals to Fe(II) bearing minerals. These reduced phases remain stationary,and react with oxidized dissolved contaminants that migrate through thetreated zone in the aquifer [215]. This approach has been demonstrated on apilot scale to treat groundwater contaminated by Cr(VI) at the Hanford Site insoutheastern Washington State, USA [215].
Biochemical or biological methods are based on the introduction or
stimulation of microorganisms that change the chemical environment [216].

38

Injection solution

Mobile field laboratory

Office storage
trailerContaminant
plume fromupgradientsourceFIG. 11. In situ redox manipulation (courtesy of Batelle).

Depending on the circumstances and intentions, a (enzymatic) reductive oroxidative precipitation of radionuclides can be effected. The application would
be similar to creating a biowall, as discussed in Section 5.5.
Thermal treatments use heat processes to immobilize the contaminant.
Thermal treatment, however, generally is not economically efficient for
dispersed radioactive contamination [5, 15].
Ex situ treatments are carried out in some sort of plant, either on or off
the site. After treatment, the material is either returned or disposed of in anengineered repository. A number of treatment techniques can be used for bothin situ and ex situ treatments, the method of application varying in each case.
Ex situ methods are discussed in Section 4.3.
Organic polymers and water glass are also used to immobilize surface
contamination. The main effect is to enhance the cohesive properties of
topsoils, thus preventing wind and water erosion [217] (Fig. 12). Depending on
the formulation, infiltration of rainwater may also be impeded and thus thedownward migration of radionuclides retarded.

39

Toxic substances

Water

Air

PEC

Soil

FIG. 12. Binding of soil particles and entrapment of contaminants using organic
polymers (S. Mikheykin, VNIICHT, Moscow, personal communication).

Over the years consultants and contractors have developed a wide rangeof proprietary engineering applications based on the fundamental processes
outlined above (e.g. Refs [218, 219]).

5.5. BIOLOGICAL BARRIER WALLS (BIOWALLS)
A biowall is an in situ barrier that relies on biological processes to restrict
the migration of radionuclides (Fig. 13). The application of the technology ismost appropriate to geological formations with significant permeability (e.g.
sands, sandstone and permeable limestones) and no preferential flowpaths
such as open cracks and fissures. A biowall can be emplaced downstream fromthe contaminated site or constructed in situ via the formation of biofilms andbiocolloids [220–226]. The development of a biowall requires the introductionof suitable microorganisms and the provision of nutrients and essential
elements to further their propagation. Adjustments to the pH or redox
potential may also be required to initiate bacterial growth.
The effectiveness of biowalls results from:40

Microorganism

Geomatrix

Borehole to inject
selected microorganisms
and nutrientsContamination

Extracellularpolymericsubstance

BiowallAquifer

FIG. 13. The principle of a biowall.

(a)

(b)

(c)

The physical reduction of permeability and hence groundwater flow by
the microbial population. This effect can be enhanced by the use of ultramicro cells (less than 100 nm). In the course of growth by metabolism
they increase in size and may completely fill the pore space.
The generation of metabolites capable of restricting the migration ofradionuclides through the barrier wall. Such metabolites are mainly
extracellular polymeric substances (EPSs), commonly termed ‘slimes’,
which the microbial cells use for attaching themselves to the substrate.
These EPSs also fill pore spaces and thus reduce permeability.
The sequestering of radionuclides from groundwaters by complexation[227–229], although it should be noted that subsequent mobilization ofthese colloidal species could constitute an additional transportmechanism [230]. Microbial action can also be utilized to modify
groundwater chemistry (e.g. sulphate reducing bacteria) to immobilize
redox sensitive species [231] such as uranium [232, 233] or to prevent theformation of acid drainage [234]. The latter two methods would act in asimilar way to an inorganic reactive wall. Several microbial strains arecommercially available.

415.6. PHYTOSTABILIZATION
Phytostabilization, the development of a stable and permanent
vegetation cover, reduces the risk of erosion of contaminated soil from sparsely
or non-vegetated land [235], thus reducing waterborne and dustborne exposure
pathways. In addition to preventing erosion, this technique may change themobility of potentially toxic elements by reducing concentrations in the soil
water and on freely exchangeable sites within the soil matrix. Both processes
alter the speciation of soil metals, reducing the potential environmental impact.
The technologies draw upon fundamental plant and soil chemical processes aswell as established agricultural practices. The development of a stable and selfperpetuating ecosystem as a result of this type of treatment may haveadditional benefits, as in some circumstances plant root activity may also
influence metal speciation by changes in redox potential or the secretion of
protons and chelating agents. The microflora associated with root systems may
also be involved in these processes. The rainwater infiltration rate is reduced byplant induced evapotranspiration, thus reducing the potential for leaching andacid drainage generation [236–238]. Some examples of phytostabilization
approaches are presented below.
Soil stabilization is very important on certain types of arable land toprevent horizontal radionuclide migration due to water and wind erosion (seeFig. 14). The 137Cs activity in topsoil in valleys may be increased by 30–80% as

Vertical migration control
Constructed protection dyke

Glubokye Lake

Cooling
pond
Projected protection dyke

Chernobyl nuclear power plant

Horizontal migration control

Sediment stabilization

FIG. 14. Phytostabilization approaches at the contaminated area of the Dnieper close
to the Chernobyl nuclear power plant.

42

a result of run-off of fine soil particles, as shown in field experiments in Belarus[142, 150]. Crop rotation with perennial grasses covering up to 50–80% of thecultivated area and avoiding row crops reduces contaminated topsoil loss from
10–20 t/ha to 2–3 t/ha. On slopes, deep soil tillage without overturning thearable layer is needed. Conventional ploughing with overturning of the arable
horizon should only be carried out to destroy and plough in old turf [239].
Good cultivation practices, such as ploughing parallel to the slope, rather thanup and down, will also reduce erosion.
Wind erosion may occur on sandy soils and on drained peaty soils. It
is recommended to eliminate root crops on soils for which soil loss amounts
to 8–15 t/ha or more [143, 240]. The major area (50–80%) of crop rotation
should be under perennial grasses. A smaller area can be allocated to winterand spring cereals and to annual grasses. In any case, soils should be undervegetation cover throughout the year, preferentially under perennial grasses. In
such a manner soil loss due to wind erosion can be reduced to 2 t/ha [239].
Remedial actions are still being investigated to control the radionuclide
efflux from the Chernobyl exclusion zone within the Dnieper catchment
system. This and adjacent drainage basins form a wide area from whichcontaminated waters flow and sediments are transported downstream throughthe Pripyat and Dnieper Rivers across Ukraine and to the Black Sea. Phytostabilization techniques could in this context also be considered as remedial
options. Three phytorehabilitation approaches involving willow plantations
have been studied [127] (Fig. 14): (1) the effect of willow plantations on verticalmigration of radionuclides; (2) the effect on the stabilization of the Chernobyl
cooling pond sediments; and (3) for lateral erosion control.
The area of interest for studying the vertical migration control by willows
was an extremely contaminated zone of 16 km2 on the left bank of the Pripyat
River (between 3.7 and 18.5 TBq/km2 90Sr and 137Cs and 0.37 TBq/km2 Pu),
which is partly protected from spring floods by a dam. Through modelling
exercises it was shown that, due to their high evapotranspiration rate, willow
SRC stands are expected to lower the groundwater table level by 100–200 cm infertilized stands. Without fertilization, a lowering of the groundwater tablelevel of less than 50 cm was predicted. Since the immobilization potential of137
Cs and 90Sr in the willow wood is limited, the influence of plant uptake on
migration remains low.
Following the closure of the Chernobyl nuclear power plant, the waterlevel of the cooling pond (22.5 km2; depths between 1.5 and 15 m; with about
111 TBq 137Cs and 37 TBq 90Sr) will drop by 4–7 m, and 15 km2 of the sediments
will become exposed and may be in need of stabilization. To this end the
SALIMAT option was investigated [241]. SALIMATs consist of a roll of willow
rods (stems) rolled around central disposable tubes that are unwound by43

dragging them across the lagoon. Small scale tests have demonstrated that
SALIMATs establish well on contaminated pond sediments and produce a fullvegetation cover during the second year. The approximated cost of the phytostabilization option ranges between €0.8 million and €1.9 million for thereclamation of 15 km2 of sediments, which is low compared with the prospected
cost of removal of the sediments ($6 million, transport and disposal costs notincluded) or maintenance of the present water level ($200 000 per year).
The project area for horizontal erosion control was on the right bank of
the Pripyat River, which was significantly less contaminated than the left bankand is not protected by a dam. After inundation, part of the activity is eroded
and transported to the Pripyat River with the withdrawing water. It wascalculated that even in the event of extremely high flooding, a dense willow
plantation will effectively decrease horizontal soil erosion and the concomitanttransport of radionuclides into the Dnieper River system.
Vegetation or revegetation is a commonly employed measure for the
capping of engineered waste disposal facilities and mining residues such as
spoil heaps [242] or tailings ponds. The final step in closing out an
impoundment for uranium mill tailings is the design and placement of a coverthat will give long term stability and control to acceptable levels radon
emanation, gamma radiation, erosion of the cover and tailings, and infiltrationand precipitation into the tailings and heaps. Surface vegetation can beeffective in protecting tailings or a tailings cover from water and wind erosion.
Factors affecting the effectiveness of surface revegetation on impoundments
can be broadly classed into climatological and agrobiological factors. Plants
should be chosen to match the local climatic conditions. From an agrobiologicalperspective, the nature of the ore and the milling process will largely determine
whether uncovered tailings are capable of sustaining growth. Considerable
efforts to improve unfavourable properties such as low or high pH values andlow plant nutrient content will usually be required before tailings can sustain
growth. Depending on the substrate, revegetation requires preparation andamelioration of the topsoil, including removal, for example, of acid generatingminerals [243, 244]. Techniques and strategies to overcome such difficulties
have been developed [245], for example hydroseeding or the use of compost
from organic household refuse [246]. The method may be limited to low
contaminant concentrations, owing to the (root) toxicity of higher concentrations. An adequate soil cover may need to be established.
Water and wind erosion are the primary concern at unvegetated tailings
or waste rock piles. While a vegetation cover will decrease erosion, it may raise
concerns that it promotes radon emanation by drying out the tailings andtailings covers. Tree roots may penetrate the contaminated material and disturb
the integrity of engineered covers. Conversely, the increased evapotranspi-

44

ration following vegetation establishment alters the water balance of a tailings
or waste rock facility and will decrease seepage. The effect of vegetation due tothe effect of plant roots on the physicochemical characteristics of tailingmaterial has not been studied very intensively so far.
At a 35 year old reclaimed site on a uranium mining dump near Schlema,
Germany, it was concluded that vegetation cover could reduce infiltration by
40–60% due to interception by the canopy (25–40%) and increased transpiration [247]. It was further found that of the 165 000 g/ha of uranium in the soil
(30 cm depth), only 4 g/ha was in the above ground plant parts and 510 g/ha in
the below ground plant parts. Since most (90%) of the uranium taken up during
the growing season is recycled (returned to the soil) with pine needles, uranium
dispersion by uptake through vegetation is minimal. It may be concluded fromthese preliminary results that forest vegetation may reduce the infiltration rate
and will disfavour radionuclide dispersion.
The proper design of tailings covers is crucial to ensure their long term
stability with respect to plant intrusion [248]. Since plant roots can penetratecompacted sealing layers (trees can have roots reaching down 3–4 m) and sincetrees need to have a certain degree of mechanical support in order to minimize
the probability of uprooting, a vegetation substrate depth of at least 1.5 m isrequired. The vegetation substrate layer must be such that the critical suction is
not exceeded at the top of the clay seal. It must be thick enough for plants tofind sufficient water and nutrients so as to prevent the generation of a highsuction at the seal. Cracks resulting from such suctions become accessible toroots and can be widened as further water is extracted.
In addition to the mechanical effects of soil stabilization and water
management, revegetation has aesthetic benefits and sometimes culturalconnotations, in particular on native or aboriginal lands [249]. The choice ofvegetation cover may also effect some sort of institutional control; for example,
converting contaminated agricultural land into forestry reserves interrupts a
potential exposure pathway via the food chain. It has to be ascertained,
however, that no other exposure pathway is opened up, for instance via burning
contaminated firewood.

5.7. CONSTRUCTED WETLANDS
Constructed wetlands are engineered, human-made ecosystems specifically designed to treat wastewater, mine drainage and other waters by
optimizing the biological, physical and chemical processes that occur in natural
wetland systems (Fig. 15). Constructed wetlands can provide an effective,
economical and environmentally sound treatment of wastewater, and serve as

45

Gravel,sand

Organic
substrate
Impermeable
liner

Inflow

Overflow

FIG. 15. Schematic cross-section of a constructed wetland.

wildlife habitats. The conceptual design leads to either the (permanent)
fixation of the contaminants in situ or to plant uptake with the view toharvesting shoots later for further treatment and disposal. The latter option is
discussed in more detail in Section 6.10, while the following largely focuses onin situ fixation.
The concept for such constructed wetlands was originally developed totreat domestic effluents for residual organic material, for example following
mechanical and biological (activated sludge) treatment steps [250], and has
found widespread application in particular for the treatment of (acid) mineeffluents [251–273]. The Chernobyl accident also prompted research in thisdirection [274]. A critical appraisal of this application is given in Ref. [275], andthe mechanisms acting are reviewed, for example, in Refs [276, 277]. The
concept was further developed to recover metals from low grade ores or minewater by a process known as phytomining (see also Section 6.10).
Constructed wetland systems are grouped into three main types [278]: (1)free water surface systems; (2) subsurface flow systems; and (3) aquatic plantsystems. Free water surface systems, or soil substrate systems, consist of aquaticplants rooted in a soil substrate within a constructed earthen basin that may or
may not be lined, depending on the soil permeability and groundwater
protection requirements [279] (Fig. 16). Free water surface systems aredesigned to accept preliminarily treated, low velocity wastewater, in plug flow,
over the top of the soil media or at a depth of between 2 and 45 cm. Subsurface
flow systems are typically gravel substrate systems that are similar to free water
surface systems; however, aquatic vegetation is planted in gravel or crushedstone and wastewater flows approximately 15 cm below the surface of themedia. The aggregate typically has a depth of between 30 and 60 cm. No visible
surface flow is evident in subsurface flow systems [279]. Aquatic plant systems
are also similar to free water surface systems, but the water is located in deeper
ponds and floating aquatic plants or submerged plants are used [281]. Whereavailable, natural ponds may be used. Where they exist, natural wetlands and

46

(a)

(b)
Pump

Spillway
Culvert

0

100 m

Inlet

FIG. 16. (a) Image of a constructed wetland; (b) plan view of a constructed wetland [280].

47

bogs can be used as traps for radionuclides and other metals [282, 283],
although this might be better classed as biosorption or natural attenuation,
since it is mostly the decaying organic matter that effects retention. Studies onnatural analogues for radionuclide migration have demonstrated thismechanism to be effective for thousands of years [114].
Early research revealed that phytoextraction via constructed wetlands(used to purify water) was ineffective because it was difficult to remove
inorganic elements that precipitated from the water into the sediments. In
addition, floating plant systems, with subsequent biomass harvesting, were
determined to be inefficient and uneconomic [284].
After three years of operating a pilot constructed wetland to treat themine water from the flooded Pöhla Tellerhäuser mine at Wismut, Germany, it
was shown that the system removed iron, arsenic, manganese and radium.
Removal processes were based on the geochemical characteristics of thecontaminants. For manganese and 226Ra, removal was also partially throughbiofilm formation. Uranium was not removed, given the high pH and thepresence of high bicarbonate concentrations [285]. It is hence clear that processeffectiveness in constructed wetlands depends on the speciation of the radionuclides concerned [286] and hence on the control of the governing parameters inthe surface and pore waters, such as pH [287], and that waters may need to be
subject to enhancement by additives [288] or pretreatment [289].
Wetlands may be constructed with the main objective of excluding
atmospheric oxygen from material that would generate acid from the oxidationof pyrite and other sulphides [290]. This method, however, is likely to beeffective only in regions where precipitation is higher than evapotranspiration(i.e. in temperate and humid tropical climates). Climatic conditions limit thegeneral applicability of wetlands. Extended periods of deep frost as well as arid
conditions are unfavourable. If, however, effluents only arise during frost free
periods, it may be possible to operate wetlands in fairly high latitudes oraltitudes.
Passive water treatment technologies such as constructed wetlands atabandoned mining sites may be appropriate for small contaminant loads.
However, long term stability and resilience with respect to external disturbances and recovery are of major concern for both wetland operators andregulators [291]. Technical guidance for designing and operating constructedwetlands may be limited, owing to a lack of long term operational data.
Potential seasonal variability and impact on wildlife may negatively affectsystem operation and securing permits, respectively [292]. Relatively large
parcels of land are required and water consumption is high, owing to large
evapotranspiration rates in some areas.

48

6. REMOVAL
6.1. PRECONDITIONS AND CONSTRAINTS
It must be noted that in general any method relying on the removal ofcontaminated soil is likely to require substitution of the removed material withclean (top)soil. Therefore, in addition to considerations with respect totechnical feasibility, an economic source of clean soil will be required to makethis option viable. Conversely, a precondition for any removal option is theavailability of a suitable disposal site for the excavated materials, whether they
are left untreated or whether they are conditioned before emplacement.

6.2. IMMOBILIZATION AND SOLIDIFICATION (EX SITU)
A wide variety of agents have been used, or proposed, for the solidification of excavated materials [212, 213, 293]. Often the objective is not only toimmobilize the contaminants but to add value to the waste material by
converting it into a useful product, for example for construction purposes. Use
in general construction as a substitute for valuable raw materials requires
special testing and licensing procedures to ensure environmental compatibility
and compliance with quality criteria such as compressive strength, freeze–thaw
cycle stability, leachability, etc. Solidified wastes may also be used in theconstruction of cappings, etc., for (hazardous waste) landfills. In cases where nofurther use is envisaged, minimization of the volume increase by the solidification agents is desirable to save valuable raw materials and repository space[213]. If only small volumes arise, the material may be combined with material
from other waste streams requiring a similar immobilization treatment.
Combining waste streams can make the process more economically viable, as
products in marketable quantities are produced.
The treatment may be undertaken on or off the site at dedicated facilities.
In the case of off-site treatment, the material has to comply with the applicabletransport regulations and must meet the appropriate safety criteria while being
handled. The additional risk from transporting material must be worked into
the respective safety and cost–benefit analyses.
As for in situ immobilization, possible (cold bonding) solidification
agents include OPC and other hydraulic binders such as certain fly ashes,
gypsum, silicate gels (water glass) [294] and organic polymers, particularly
acrylic and epoxy resins. The overall cost of treatment obviously depends on
the amounts or volumes requiring treatment, and some binders are more49

economic in larger quantities than others. Organic polymers, in most cases,
appear to be applicable only to smaller volumes of material. The conversion of
gypsum (CaSO4·2H2O) into a semi-hydrate (plaster of Paris, CaSO4·½H2O)
requires heat treatment in a kiln. It is only the plaster, not the gypsum, thatexhibits the curing properties required of binders.
The main conclusion of a recent report [295] on the European perspectiveof NORM waste treatment was that immobilization is not widely used or
accepted as a treatment. Many companies regard this type of technology as lessfeasible for NORM waste material and hence have not pursued thedevelopment of immobilization techniques as a waste treatment process.
However, for treating the radioactive remainder of a separation step, immobilization is widely seen as a treatment with a high potential.
Into this latter classification would also fall ground freezing as atemporary measure to prevent the dispersal of contaminants [296, 297]. Either
an impermeable screen around a contamination can be established or thecontaminated material itself can be frozen in order to facilitate its handling. In
either case, it is unlikely that in the present context of low level dispersed
contamination this method would find a field of application.
The removal of a contaminated topsoil layer is, of course, the mosteffective measure, but generates large quantities of waste and is only applicableto small areas of land [138]. Moreover, the most fertile layer of the soil isremoved in the process. The overall efficiency of such a measure depends verymuch on the operating conditions and on the distribution of the contaminationin the vicinity of a critical group [151]. In Belarus, the Russian Federation andUkraine the removal of contaminated topsoil was recommended for all
settlements where the 137Cs activity exceeded 555 kBq/m2 and for 25–33% ofsettlements where the 137Cs activity was in the range of 370 to 555 kBq/m2. It
was estimated that it would incur costs of about €325 per inhabitant [151].

6.3. SURFACE AND GROUNDWATERS: PUMP AND TREAT
Pump and treat systems (Fig. 17) were the baseline for remediating
groundwaters throughout the 1990s [10, 298–301]. The technology is based on
the assumption that contaminant concentrations can be reduced or removed by
employing ion exchange or sorption [302] and precipitation processes [303].
Some attempts have been made to use electrolysis [304] or (reverse) osmosis inpump and treat systems. Chemicals have also been added underground in an
attempt to enhance recovery rates [305].
A National Academy of Sciences report [306] provides a comprehensive
assessment of the effectiveness of pump and treat systems for the remediation

50

Contaminants

Abstraction well

Conditioning
disposal

Injection well

Treatment
plant

Contaminant plume

Aquifer

FIG. 17. A pump and treat system.

of subsurface contamination. The report found that pump and treat isinefficient as a source removal technology, although it can reduce source term
volumes. In line with other methods based on changing the distribution
between two different phases of a contaminant, this method becomes increasingly inefficient as the concentration gradient between, for example, species
sorbed on the solid matrix and aqueous species diminishes. Large quantities ofgroundwater may have to be pumped and treated to remove only small
amounts of contaminant. Removal in situ is inefficient, owing to tailing or masstransfer limitations. A further complication arises from the fact that not all porewater is mobile. Contaminants may be trapped in dead end pores and released
into the mobile pore water only by diffusive processes, which is one of themechanisms responsible for the tailing. Although various configurations ofabstraction wells, etc., have been investigated with a view to increasing thedegree of hydraulic connectedness and hence efficiency [307, 308], these configurations have been unable to overcome the fundamental constraints on diffusion.

51

Undesirable water properties, for example low pH values, as is often thecase with mine effluents or disposal facility leachates, may pose special
problems during processing [309]; a neutralization step might be required [310].
For these reasons it is unlikely that simple pump and treat methods will
have much scope for application in situations with relatively low levels ofcontamination.

6.4. ENHANCED RECOVERY
It may be desirable to chemically treat aquifers in order to facilitate or
accelerate the recovery of radionuclides or to lower residual concentrations inpump and treat scenarios [305]. Such methods are often termed ‘soil flushing’[311]. After removal of the contaminant and before being reinjected, thepumped water is dosed with lixiviants, for example acid, surfactants [312],
complexing agents such as EDTA [313, 314] or inert electrolytes, to replacesorbed radionuclides. However, unwanted side effects, such as dissolution of
the rock matrix, may be difficult to predict. Some of the available extraction
methods are used in hydrometallurgy to enhance metal value recovery [315,316]. Figure 18 shows the principal layout for the treatment of an aquifer, while
Fig. 19 shows the arrangement for treating the unsaturated zone above anaquifer.
Biological in situ leaching may be appropriate as a technique forenhanced recovery. More details on this method are given in Section 6.9.
Electrochemical methods for enhancing recovery of radionuclides inaqueous solutions have been proposed [304, 317–319]. If an electric field isapplied to a solution, inorganic and organic ions migrate according to theircharges to the respective electrodes (Fig. 20). Two primary mechanisms
transport contaminants through the soil towards one or the other electrode:electromigration and electro-osmosis. In electromigration, charged particles
are transported through the substrate. Electrolysis arrangements concentrate
metal ions on the cathode and can aid the oxidation of organic contaminants. In
contrast, electro-osmosis is the movement of a liquid containing ions relative toa stationary charged surface. The direction and rate of movement of an ionicspecies will depend on its charge, both in magnitude and polarity, as well as on
the magnitude of the electro-osmosis induced flow velocity. Non-ionic species,
both inorganic and organic, will also be transported along with the electroosmosis induced water flow.
Different types of electrode material have been tested to improve
performance, including porous ceramics and the rather novel carbon aerogels

52

Excess water discharge
Treatment
plant
Abstraction well

Abstraction well

Lixiviant added
Injection well

Contaminant plume

Groundwater
flow

FIG. 18. Sketch of an in situ leaching or enhanced recovery arrangement.

Excess water discharge

Treatment
plant

Lixiviant
addition

Abstraction well
SprinklingsystemContaminanted soil

Groundwater
flow

FIG. 19. Sprinkling of contaminated soil in the vadose zone to remove contamination.

53

AnodeCathode=Contaminant plume

FIG. 20. Generic layout for remediation by electrolysis and electro-osmosis.

that increase the effective surface area [302]. Electro-osmosis may be
combined with other techniques to remove contaminants from low permeability geomatrices such as clays. LASAGNA is a technology demonstration
project designed to evaluate a combination of techniques [168].

6.5. CHEMICAL EXTRACTION (EX SITU)
Disposal of contaminated soil is likely to be an inefficient option in mostcases, owing to the large volumes involved [24], even though compaction may
help to reduce the volume [320]. Separation or extraction of the contaminants
is preferable where no suitable in situ technique is available, and a variety ofproprietary processes have been developed over the past decade or so (e.g.
Ref. [321]). Excavated soils can be treated with inorganic and organic solvents
to dissolve and remove metals, including radionuclides [322]. The effectivenessand efficiency of a given solvent will depend on the type of binding between theradionuclide and the soil substrate and on the chemical species of the radionuclide. The choice of chemicals that can be applied is much more varied than forin situ treatment, given the better control of the processes and the fact that theoperation can be carried out in closed reaction vessels. Considerations ofenvironmental impact from the remediation operation, for example unwanted54effects on the groundwater and aquifers, are restricted to considerations that
apply to similar industrial operations.
The solvents used can be either salt solutions that aim to replace radionuclides from sorption sites, as in backwashing ion exchangers, or acids that
destroy substrates, such as carbonates, chelating agents [323–326] or surfactants[312] that bring sorbed ions into solution, or organic solvents that solubilize
neutral species.
The recovery of low concentrations of uranium and other radionuclides
from the process solutions can be a problem in itself. It is typically carried out
by ion exchange. Suitable exchangers are synthetic organic polymers [327] orcertain mineral species [328], such as zeolites.
Physicochemical methods, such as dialysis or membrane separation [329],
sometimes in conjunction with electrochemical methods, have also been
applied to remove radionuclides from soils.
Treatment residues themselves, such as sewage sludges, may need further
treatment to remove radionuclides before they can be utilized or disposed of[330].
While being offered by a wide range of commercial contractors, it must be
noted that, typically, extraction or soil washing processes destroy the soil’sfunctionality and result in a sterile product. When reuse as topsoil is planned,
addition of humus formers, for example compost, is necessary.

6.6. HYDROMETALLURGICAL METHODS
Ex situ methods to remove metals from soils and rocks, such as heap
leaching, have undergone considerable development in the field of hydrometallurgy and solution mining [316, 331–333] and have been adapted to
remediation problems [315, 334, 335]. Although the (bio)chemical processes
are similar to those discussed in Sections 6.5 and 6.9, these methods arepresented here separately because of their industrial application background.
Large quantities of material are excavated and piled on specially prepared pads
or polders. The leaching solution is sprayed or trickled over these ‘heaps’ atintervals and the leachate is collected for further processing (Fig. 21). The
processing follows similar lines as that for treating pumped groundwaters.

55

Sprinkling system
Leaching
solution

Leachate

Retainingdam

Drainage/leachatecollecting system

Bottomliner

FIG. 21. A heap leaching system.

6.7. SEGREGATION
Contamination is often associated with particular size fractions or
mineral phases of a soil. Separation and segregation of the contaminatedfraction will greatly reduce the amount of material requiring further treatment
and disposal, while freeing the reminder for reuse. Extraction processes cantherefore usually be made more efficient by a prior size fractionation of thegeological materials [336–338], since contaminants tend to accumulate in theclay fraction because of the larger number of sorption sites on clay minerals
and their often higher specific binding energy. Radionuclides and other heavy
metals may also substitute for common ions of similar ionic radius and charge
as counter ions in interlayers. In certain cases neoformation of clay minerals
can incorporate radionuclides in the mineral lattice.
A variety of separation techniques have been borrowed from mineralprocessing, including mechanical sieving and screening, hydraulic size fractionation in, for example, settling tanks or hydrocyclones, specific gravity
separators such as shaking tables or sluices, surface chemistry related processes
such as froth floatation, and processes based on the different magnetic susceptibilities of different minerals. A combination of these techniques may berequired to isolate the relevant fractions. Segregation is often the first step
before one of the above chemical extraction methods is applied. The latter are
also referred to as soil washing, if they form part of an extraction procedure.

6.8. BIOSORPTION
Scientific and technological developments over the past few decades havemade possible the application of biological processes to the solution of complex

56

environmental problems (Refs [339–341] and the proceedings of the international biohydrometallurgy symposia). Microorganisms, that is to say single cellorganisms such as bacteria or fungi, have been used as minute biologicalreactors that can efficiently and economically carry out specialized operations.
Microbial biomass, whether living or not, has been shown to selectively
sequester and retain elements from dilute aqueous solutions via a process
named biosorption.
Through the process of biosorption the biosorbed species are selectively
removed from the solution and are retained inside the microbial cells (biomass)in concentrations that are several orders of magnitude higher than those in theoriginal solution. Heavy metals and radionuclides are taken up into cellularcomponents such as the cell walls of certain microorganisms, which then can be
harvested, carrying along the sequestered radionuclides. Biosorption is being
explored in hydrometallurgy to concentrate metal bearing solutions, forexample from heap leaching, and in the treatment of contaminated mining
effluents [291, 316, 342, 343].
Engineering developments in the area of biosorption have led to thedesign of engineered biosorbents, microbial biomass cells or cellularcomponents immobilized on or within various matrices, thus acquiring the formof small particles such as those of conventional adsorbents (e.g. activatedcarbon) or ion exchange resins. Organic cellular material derived from higher
plants or algae have also been proposed as the basic material for manufacturingbiosorbents that can be used for the extraction of metals, including radionuclides [344–354].
Biosorption methods are largely ex situ methods applicable for diluting
contaminated solutions such as groundwaters or seepage. The contaminatedsolution is pumped into engineered reactors, in which it contacts theimmobilized microbial biomass under optimized conditions (solution pH, flow
rate, etc.). The contaminants are retained in an insoluble form by the biomass
and the treated solution is let out of the reactor. The process of biosorption isreversible under certain conditions, which means that after the biosorbent isexhausted it could potentially be used for regeneration, releasing thepreviously held radionuclides in a small volume of the regenerating solution.
Alternatively the biosorbent could be used once through and then disposed ofappropriately.
Biosorption is an equilibrium process, with solution pH playing the roleof the master variable, since it defines the speciation of the elements in thesolution. This also means that the key driving force that dictates the biosorptive
uptake capacity of the biomass in terms of mass of biosorbed species per unit
mass of biosorbent (also referred to as the loading capacity) is the residual57

concentration of the contaminants after treatment and not the initial
contaminant concentration [355].
The optimal biosorption pH depends on the biomass used and on theelements being removed; for example, the biosorption of uranium by thefungus Rhizopus arrhizus appears to be optimal at pH4, with significant
reduction of the metal uptake capacity as the pH drops to pH2. The increased
concentration of hydrogen ions at the acidic pH along with the chemical effects
on the cell walls of the microorganisms are responsible for this reduction incapacity [356, 357]. However, increasing the pH towards neutral values may
again create operational problems, depending on the composition of thecontact solution. The hydrolysis and subsequent precipitation of ferric ionswhich adsorb on to (coat) the biosorbent adversely affect the biosorption
process [227].
The biosorption of metals by algal biomass is another example in which
the sequestering of metals such as lead, zinc or copper by microorganisms suchas Chlorella vulgaris, Chlorella regularis or Chlamydomonas sp. is optimal in
the range of pH6 to pH8. The biosorption of oxyanions such as chromates orselenates by the same type of algae has an optimal biosorption pH in the acidic
range of pH2 to pH3 [358, 359].
Biosorption of 226Ra by several types of microorganisms, such as
Rhizopus arrhizus, Aspergillus niger and Streptomyces niveus, exhibited an
optimal contact pH in the neutral to alkaline range, with corresponding radium
equilibrium uptake capacities in the range of tens of MBq/g. It is therefore
obvious that optimization of biosorption processes should be made on a case bycase basis and requires increased care so that the process will perform satisfactorily.
Considerable efforts have been made to understand the underlying
mechanisms of biosorption and to improve the process efficiency. The available
information has shown that cell walls are the major biosorption functional sites
for heavy metals, uranium and thorium. It has also been shown that EPSs playa significant role in biosorption [360]. The molecular level understanding of thebiosorptive processes is still limited to selected pairs of metals and microorganisms. The microbial biomass provides ligand groups on to which the metal
species bind. In addition, sorptive and hydrolysis processes play a role. Three
major classes of microbial biopolymers (proteins, nucleic acids and polysaccharides) provide biosorption sites. Different ionic species of a given element
might exhibit preference for a different binding site. Should the preference of
one metal ion for a ligand be similar to that of another ion, a biosorptioncompetition effect might be observed if both elements are simultaneously
present in the contact solution.

58

A model of bisorption competition effects that is based on Pearson’sclassification of metals [361] has been reported as a basic tool for understanding such effects. On the basis of this model, significant ionic competition
effects can be observed for metals belonging to the same Pearson classification
class. Elements belonging to different classes demonstrate limited competition,
while elements belonging to the Pearson’s classification borderline class areaffected by the presence of competing co-ions [362]. Additional systematic
work for the mechanistic understanding of biosorptive processes and theassociated ionic competition effects is required.
Numerical simulation techniques play an important role in designing andassessing remediation processes, including those using biotechnologicalmethods [363–365]. Although biosorption using inactive microbial biomass has
been demonstrated to be effective in substantially removing (and in some cases
recovering) targeted radionuclides such as uranium, radium and thorium fromcontaminated solutions, a full scale commercial application is not yet available.
The use of living microorganisms in innovative reactor configurations has
recently been under investigation for the same purposes as conventional
biosorption. This approach to the biological sequestering of metals has substantially different requirements and operating conditions than conventional
inactive biomass biosorption. This alternative biotechnological approach isoften referred to as bioaccumulation or bioprecipitation and is showing
excellent potential.

6.9. BIOLEACHING
Bioleaching occurs naturally when microorganisms assist in the slow
weathering of out-cropping sulphide ore bodies. Bioleaching is an establishedbiotechnological process for the dissolution and hence mobilization of valuable
metals from ores by microorganisms [366–371]. Metals for which this technique
is mainly employed are copper, cobalt, nickel, zinc, gold, silver and uranium. It
is estimated that about 20–30% of the world’s copper production originatesfrom bioleaching; in the case of uranium it is judged to be about 5–10%.
Bioleaching has also been promoted as a cost efficient method for metal value
recovery in developing countries, and its applicability in this context has
recently been reviewed [372].
Bioleaching also has scope for application in reworking waste material
from mining for enhanced recovery of metals, including radionuclides, which
has the potential to reduce the environmental burden. The method has been
explicitly applied to the remediation of uranium and other mining legacies[373, 374]. The pathways of the resulting contaminated waters have to be59carefully controlled, for example by arrangements similar to those for heap
leaching (Fig. 21). Microbially mediated leaching processes frequently have theunwanted side effect of AMD generation, for example by pyrite oxidation[375].
The types of ore that are amenable to bioleaching comprise sulphides,
carbonates and oxides. The groups of microorganisms involved are mainlybacteria and fungi. In some cases algae and lichens may also play a role.
Various mechanisms are involved, depending on the type of ore in question.
In the case of sulphidic minerals the predominant dissolution causing
microorganisms are acidophilic (meaning organisms living between pH0 andpH5) bacteria of the sulphur and iron cycles, namely Acidithiobacillus (abbreviation A., former name Thiobacillus) ferrooxidans, A. thiooxidans,
Leptospirillum (abbreviation L.) ferrooxidans, A. caldus, Metallogenium sp.,
Sulfobacillus thermosulfidooxidans, Sulfolobus sp., Acidianus brierleyi andseveral others. The species of Acidithiobacillus live in the moderatetemperature range (0–45ºC), Metallogenium and Sulfobacillus thrive atelevated temperatures (40–65ºC) and Sulfolobus and Acidianus arethermophiles growing from 65 to 90ºC.
The dissolution is generally effected by two mechanisms, depending on
the type of mineral to be dissolved:
(a)

(b)

Pyrite and molybdenite and a few other minerals of the same structure
can only be dissolved by an oxidizing attack on their crystal lattice, owing
to their electronic configuration (non-bonding outer orbitals) [376]. The
bacteria able to do this are the Fe(II) oxidizing A. ferrooxidans, L.
ferrooxidans and Acidianus sp. This mechanism is known as the thiosulphate mechanism.
All other sulphidic minerals possess bonding outer orbitals and thus are
more or less dissolvable by a hydrolytic attack involving protons. In
addition, Fe(III) ions further the dissolution by an oxidizing attack. Theseminerals may consequently be dissolved by all the above mentioned
bacteria of the sulphur and iron cycles. This dissolution process is known
as the polysulphide mechanism.

In both cases, the dissolution of the mineral is mainly effected by bacteria
attached to the surface of the respective mineral. The compounds mediatingsuch attachment are EPSs (‘slimes’). The EPSs consist, from a chemical pointof view, mainly of lipids, carbohydrates, sugar acids (uronic acids) andcomplexed, inorganic ions such as Fe(III) ions. The distance between thebacterial cell and the mineral substrate surface is of the order of 20 to 50 nm.
This space is filled with the EPS, creating a reaction space with unknown

60

conditions of pH, redox and ion concentrations; the reaction space is anextension of the radius of action of the cell, thus allowing it to augment its foodsupply. As a consequence, biological leaching becomes considerably
accelerated (sometimes more than 100-fold) compared with the purely
chemical process utilizing Fe(III) ions and/or protons only. In the latter process
the freely suspended planktonic cells also have to be considered, since theireffect is mainly the reoxidation of the iron ions in solution. Bioleaching is thus
an interface process and belongs to the area of nanobiotechnology.
Final products of dissolution are metal cations, Fe(III) ions, sulphate and/or sulphuric acid. The energy of the oxidation is partially conserved by thebacteria for metabolic purposes and growth. The bacteria possess specialized
cell components allowing them to conserve some of the energy in a utilizable
form (ATP, NADH). Furthermore, they need only carbon dioxide from the air
to build up their cell mass and inorganic trace elements. These are therefore
very specialized organisms; this type of metabolism is called lithoautotrophy.
The above mentioned bacteria are generally not important for carbonate
and/or oxide ores. Bacteria and fungi are used for dissolving such minerals,
which, due to an unbalanced metabolism, excrete organic acids. This requires anample supply of exogenous carbon sources, which they metabolize, and as aconsequence of either too much substrate, or a lack of essential nutrients or traceelements such as nitrogenous compounds or minerals, excrete partly in an intermediate oxidation state. Excreted acids are, for example, citric, oxalic, succinic,
malic, acetic and/or formic acid, and sugar (uronic acids) or amino acids. These
acids dissolve and/or complex metal cations and thus solubilize them.
The bioleaching technique is employed in several forms.
(a)In the case of low grade ores that for economic reasons cannot beprocessed by conventional roasting or other similar processes, a heap
leaching process is applied (Fig. 21). In the majority of cases in which thistechnique has been applied to date, the ore contained copper, zinc andtrace elements. A limited number of experiments of this type have been
performed for extracting uranium from low grade ores. One experiment
was carried out near Ronneburg, Germany, by Wismut in the 1980s,
another one at Elliott Lake in Canada. For this purpose large amounts oflow grade ores are placed on leach pads (plastic liners) or dumped invalleys with a known and impermeable geological strata and sprinkled
regularly with acidified bacteria-containing solution (which originatesfrom similar operations or from acid mine waters). The dissolved metals
and sulphate plus sulphuric acid are left to accumulate to a concentration
at which extraction processes such as solvent extraction, ion exchange
and/or electrowinning become viable. Residence times for such

61

(b)

(c)

operations range from several months to a few years. If these heaps areconstructed without consideration of the underlying geology and/or
abandoned without care, acid mine–rock drainage (ARD or AMD) may
result. Abandoned mines, mine shafts, open pits, etc., might also produce
and release ARD. Owing to the acidity combined with dissolved heavy
metals, this might result in serious environmental damage and/or evencreate new ecosystems (as in Rio Tinto, Spain).
Bioleaching has been employed experimentally at the field scale inGermany to treat heavy metal contaminated sediments [377]. The genericscheme from the raw sediment to a viable soil substrate is illustrated inFig. 22.
In the case of sulphidic concentrates, bioleaching is increasingly used forextracting precious metals. In recent years several plants have gone into
operation that use acidophilic leaching bacteria for extracting gold, nickel
and cobalt. The operation usually consists of stirred tanks (bioreactors)
with volumes of up to 1000 m3 in continuous operation. Residence times
are in the range of 3 to 7 days.

Another form of application is in situ bioleaching (see Fig. 18) [378]. This
technique is applied: (1) to treat ore bodies that are located too deep
underground for conventional mining; (2) to treat leftover ore from othermining activities, as was the case with the uranium bioleaching at Elliott Lake,
Canada; and (3) to avoid damage to the surface environment. Consequently,

Raw
sediment

GradingFinefraction

Conditioningby plants

Compostingof plants

Process water
treatment

Process
water

Bioleaching

RevitalizationSandfraction
Reagents

Lime

Soil
Metal sludge

FIG. 22. The experimental process from raw contaminated sediment to reconditioned
soil, after Ref. [377].

62

for bioleaching the host rocks are typically prepared by blasting to perforate
them and by the installation of pumping in and pumping out pipe systems forthe application, recirculation and extraction of leach solutions. The pregnant
solution is then processed in an extraction plant to extract the valuable metals.
Depending on the geological setting, such in situ leaching may pose different
hazards to the environment. Sometimes, losses of the leach solution that enter
unnoticed underlying aquifers via faults or fissures in the separating geologicalstrata are encountered.
In the case of radioactive minerals, there may also be another, unwanted
effect: an enhanced emission of radon. Comparison of the radon emission rates
and bioleaching activity at the above mentioned leaching waste heaps near
Ronneburg, Germany, has shown that high cell numbers of leaching bacteria
were found at sites with high radon emissions, whereas at sites with low
emissions only low cell numbers occurred. An explanation for this effect comesfrom the mineralogy of the ore. At Ronneburg the uranium is embedded inpyrite. Once this pyrite has been attacked by bioleaching, radon is liberated
and may escape into the atmosphere. This causes an additional exposure for the
local population and requires measures to reduce or even inhibit the biologicalprocess.

6.10. PHYTOEXTRACTION
6.10.1. Overview
The use of plants to remove contaminants from the environment andconcentrate them in above ground plant tissue is known as phytoextraction[379–394]. Phytoextraction requires that the target metal (radionuclide) be
available to the plant root, absorbed by the root and translocated from the rootto the shoot; biomass production should be substantial. The metal (radionuclide) is removed from the site by harvesting the biomass, after which it isprocessed either to recover the metal or further concentrate the metal (by athermal, microbial or chemical treatment) to facilitate disposal.
Research and development efforts have focused on two areas: (1)remediation of contaminants such as lead [395–397], arsenic, chromium,mercury and radionuclides [129, 386, 398–406]; and (2) mining, or recovery, of
inorganic compounds, mainly nickel and copper, having intrinsic economic
value.
Successful implementation of phytoextraction depends on [284]:(a)

The bioavailability of the contaminant in the environmental matrix;

63

(b)
(c)
(d)

Root uptake;
Internal translocation of the plant;
Plant tolerance.

Plant productivity (i.e. the amount of dry matter that is harvestable eachseason) and the accumulation factor (the ratio of metal in plant tissue to that in
the soil) are important design parameters [406–408]. This is clearly exemplified
by the following set of equations and tables [406]. The percentage yearly
reduction in soil activity can be calculated as:
Annual removal (%) = TF ¥ yield ¥ 100
Wsoil

(1)

where TF is the transfer factor or bioaccumulation factor (TF = Cplant/Csoil, with
Cplant the concentration of the radiocontaminant in the plant (Bq/g) and Csoil
the concentration of the contaminant in the soil) and Wsoil the weight of thecontaminated soil layer (kg/ha). As is evident from Eq. (1), the annual removal
percentage increases with yield and TF. However, TF and yield values are notindependent: a high yield is often associated with lower TFs because of growth
dilution effects.
Phytoextraction typically requires several years of operation, and thefuture trend in radionuclide concentration in the soil can be calculated from:È Ê TF ¥ yield 0.69 ˆ ˘¥ t˙
C soil, t = C soil, t = 0 exp Í- Á+
t 1 / 2 ˜¯ ˙˚ÍÎ Ë Wsoil(2)

The second term in the exponent of Eq. (2) accounts for radioactive
decay (t1/2 is the half-life of the radionuclide). For some radionuclides with long
half-lives (e.g. t1/2 for 238U is 4.5 × 109 a), this component will not affect thephytoextraction potential. For others, for example 137Cs and 90Sr, with half-livesof 30 years, the phytoextraction potential will be affected; that is, a yearly lossof 2.33% in activity occurs merely through radioactive decay. Equation (2)assumes a constant bioavailability of the contaminant (i.e. a constant TF).
Table 4 shows, for a calculated example, the percentage annual removal
per hectare for different crop yields and TFs, based on a 10 cm deep soil layerthat has a mass of 1500 t for a soil density of 1.5 kg/dm3. It should be borne inmind that if the contamination is spread to a depth of 20–50 cm in the soil,
annual removal with the biomass is reduced by a factor of 2 to 5, respectively,
compared with the figures presented in Table 6.

64

TABLE 4. PERCENTAGE YEARLY REDUCTION OF SOIL CONTAMINATION DUE TO PHYTOEXTRACTION AND RADIOACTIVE
DECAY [406]Annual reduction due tophytoextraction (%)

Annual reduction due tophytoextraction and decay (%)

Yield (t/ha)

Yield (t/ha)

TF (g/g)
5

10

15

20

30

5

10

15

20

30

0.01

0.003

0.007

0.01

0.013

0.02

2.33

2.34

2.34

2.34

2.35

0.1

0.033

0.067

0.1

0.133

0.2

2.36

2.40

2.43

2.46

2.53

1

0.33

0.67

1.00

1.33

2.00

2.66

3

3.33

3.66

4.33

2

0.67

1.33

2.00

2.67

4.00

3

3.66

4.33

5

6.33

5

1.67

3.33

5.00

6.67

10.00

4

5.66

7.33

9

12.33

10

3.33

6.67

10.00

13.33

20.00

5.66

9

12.33

15.7

22.33

Note: t½: 30 a; soil depth: 10 cm; soil density: 1.5 kg/dm3.

Yields of more than 20 t/ha and TFs higher than 0.1 (Table 4) may be
regarded as upper limits, except for strontium. This would result in an annual
reduction percentage of 0.1% (decay excluded). When TF equals 1, the annual
reduction is about 1%. Table 5 gives some ranges for TFs for the natural radionuclides uranium and radium, predominant contaminants in the NORM
industries, and the long lived fission products 137Cs and 90Sr.
By rearranging Eq. (2) the number of years needed to attain the required
reduction factor as a function of annual removal percentage can be calculated.
Table 6 presents the number of years required to attain a reduction of thecontaminant concentration up to a factor of 100, given an annual extraction

TABLE 5. RANGES FOR TFs (RATIO) BASED ON DATA FROM REFS
[122, 123, 141, 146, 402, 403, 409–416]
Total range(Bq/g plant to Bq/g soil)

Comment on conditions for upper limit

Cs

0.00025 −7.5

Brassica, organic soil

Sr0.0051 − 22

Green vegetables, sandy soil

U

0.000006 −21.13

Tubers, sandy soil

Ra

0.00029 − 0.21

Grass, sandy soil

65

TABLE 6. CALCULATED NUMBER OF YEARS REQUIRED TO
DECONTAMINATE A SOIL FOR A REQUIRED (DESIRED)
REDUCTION FACTOR AND A GIVEN ANNUAL REMOVALPERCENTAGE
Desiredreduction
factorActivity remaining,
Csoil,t /Csoil,t=0 (%)

Annual removal (%/a)
20

15

10

5

3

2

1

0.1

5

20

7

10

15

31

53

80

160

1650

10

10

10

14

22

45

76

114

229

2301

20

5

13

18

28

58

98

148

298

2994

50

2

18

27

37

76

128

194

389

3910

100

1

21

28

44

90

151

228

458

4603

3

Note: Soil depth: 10 cm; soil density: 1.5 kg/dm .

percentage or percentage reduction in radionuclide activity varying between
0.1% and 20%. With an annual removal of 0.1% it would take more than 2000
years to decontaminate a soil to 10% of its initial activity; with an annual
removal of 1%, more than 200 years are required. It is hence clear thatmeasures would need to be taken to increase the annual removal efficiency
through crop selection, or to increase the bioavailability by applying soil
additives and through technical measures (e.g. decreasing the tilled soil depth).
In most cases one has limited control over the depth of the contamination, although it may be feasible and advantageous to excavate and pile thesoil to the desired soil depth for phytoremediation purposes. One possibility is
to excavate the soil and spread it on geomembranes, which impedes roots from
penetrating to deeper layers. These membranes will also limit contaminantdispersal to the underlying clean soil, but a substantial area may be needed fortreatment. Decreasing the tilled soil depth increases the removal percentage
according to Eq. (1) and may intensify root–soil contact, and may result in an
increased TF.
The other factors influencing radionuclide bioavailability, such as crop
selection and measures to increase the bioavailability of the radionuclide ofconcern, are generally radionuclide specific. To maximize the metal content in
the biomass, it is necessary to use a combination of improved soil management
measures, for example optimizing the soil pH and mineral nutrient contents, orthe addition of agents that increase the availability of metals.
Apart from the application of soil additives to increase export with theplant biomass, plant selection may also be important for improving the66phytoextraction potential [417]. As already mentioned, there is a significant
interspecies variability in TFs (Table 9). Since the values are seldom obtained
for similar soil and growth conditions, the effect of plant species on the TFs
cannot be unambiguously derived. Observed differences between plantvarieties or cultivars have been up to a factor of 2 [403].
Improved genotypes with optimized metal uptake, translocation andtolerance, and improved biomass yield, may also be an approach to improved
phytoextraction. Plant breeding [418] and genetic engineering may open
further alleys to develop hyperaccumulating plants [419–421], but actual
research and technology development is mostly limited to heavy metals.
Although positive effects have been obtained following applications ofsoil amendments that increase element bioavailability, the effect of continuous
treatment on soil quality, plant growth and bioaccumulation is not clear. Therealso remains the question of long term effectiveness: will TFs remain constantor will they decrease as radionuclide concentrations decrease.
Effective extraction of radionuclides and heavy metals by hyperaccumulators is limited to shallow soil depths of up to 30 cm. If a contamination isfound at substantially greater depths (e.g. 2–3 m), deep rooting perennial crops
could in principle be employed, but the fraction of their roots exploring thecontaminated zone would be small and hence also the phytoextraction
potential.
There are concerns that contaminated leaf litter and associated toxic
residues [422] may result in uncontrolled dispersion of the contaminants.
Finding a safe use or disposal route for contaminated biomass will be a major
element in developing a phytoextraction scheme.
Little is known about the economics of phytoextraction, which not only
depends on the extraction efficiency but also on the costs associated with crop
management (i.e. soil management, sowing or planting (yearly returns forannual crops), harvesting, post-harvest biomass transport, biomass treatment,
potential disposal costs and site monitoring). The treatment of 1 m3 of contaminated soil (10 m2 for a 1 dm soil layer) will result in about 10–20 kg of biomass(~2–4 kg of ash) annually.
6.10.2. Uranium removalFree UO2+
2 is the uranium species most readily taken up and translocatedby plants. Since this uranium species is only present at a pH of pH5.5 or less,acidification of uranium contaminated soils may be necessary for phytoextraction [401]. The uranyl cation also binds to the soil solids and organic matter,
reducing the extent of plant uptake [423, 424]. Therefore, in addition to acidification, soil amendments that increase the availability of uranium by complex-

67

ation may also be required. In testing the role of acidification and chelating
agents on the solubilization of uranium it was found [402, 403] that, of theorganic acids and chelating agents tested, citric acid was the most effective forincreasing uranium in the soil solution. Following citric acid treatment(20 mmol/kg) the uranium accumulation in Indian mustard (Brassica juncea)
was increased 1000-fold [403] and in beet (Beta vulgaris) tenfold [402, 403].
Similar results were obtained when testing the potential for phytoextraction of uranium from a low level contaminated sandy soil using rye grass(Lolium perenne cv. Melvina), Indian mustard (Brassica juncea cv. Vitasso) andredroot pigweed (Amarathus retroflexus) [129]. The annual removal of the soil
activity with the biomass was less than 0.1%. Addition of citric acid increased
uranium uptake up to 500-fold, and extraction percentages of 2–5% appear
achievable. Citric acid addition, however, resulted in a decreased dry weight
production (all plants tested) and even plant death and crop regrowth (in thecase of rye grass). Depending on the desired contamination reduction factor(e.g. 5–50), it would still take between 30 and 200 years for the target to be met(Table 6).
6.10.3. Strontium removal
Table 7 shows the annual crop removal of 137Cs and 90Sr. It is clear fromthis table that in normal agricultural systems the annual caesium flux is small
compared with the reservoir present in the soil. The 137Cs removal rates are allless than 1%, and the highest removal is found for grassland. The high sorptionof 137Cs in soil and the typical potassium levels in soil required for optimal plant
growth all limit removal rates.
The removal of 90Sr with biomass is higher than that of 137Cs because the90
Sr availability is typically tenfold above that of caesium. The TFs of 90Sr ingreen vegetables and Brassica plants are typically around unity and the upper
levels are around 10 [409]. Phytoextraction of 90Sr has not yet been investigatedat the field scale. The high removal rates in agricultural crops (Table 7) suggestthat phytoextraction may be worth while to explore.
The highest transfers of 90Sr are typical for leguminous perennial grasses(Trifolium family) and Brassica plants [25]. Field experiments in Belarus werecarried out at the Belarussian Research Institute for Soil Science andAgrochemistry (BRISSA) [425] on light-textured soil contaminated with 90Sr
(Table 8). It was found that cow clover (Trifolium pratense) has annual green
mass yields of up to 65–75 t/ha (6–7 t/ha dry mass). The 90Sr removal values werein the range 2.5–3.6% of the total radionuclide reservoir in the soil. A change ofsoil pH from neutral (pH6.8) to moderately acid (pH4.9) enhanced the 90Sr
transfer by a factor of almost 2, but the yield of clover was reduced, so the total

68

TABLE 7. ANNUAL REMOVAL BY CROP BIOMASS OF 137Cs AND 90Sr
FOR SOME AGRICULTURAL CROPS, EXPRESSED AS A FRACTION
OF TOTAL CONTENT IN THE TILLED LAYER (ARABLE CROPS) OR
IN THE 0–12.5 cm LAYER (GRASSLAND)
Strontium crop
Caesium crop
Yield (dryStrontium TF
Caesium TF
removal (% ofremoval (% ofmatter)(g/g)(g/g)
total in soil)
total in soil)(t/ha)
Cereals(grain)5–70.0004–0.25

0.0005–0.06

0.02–0.94

0.0037–0.22

Potato tuber

6–10

0.003–0.89

0.0006–0.3

0.03–1.4

0.006–0.5

Leafy
vegetables

5–10

0.008–1.7

0.001–0.6

0.45–9.1

0.07–3.0

Grassland

10–15

0.01–1.0

0.007–1.0——

Note: The TF ranges were derived from Ref. [409].

accumulation of radionuclide per unit area was increased only by a factor of
1.5. It should be noted that when the clover is used as animal fodder, thegreater part of 90Sr activity will end up in dung and in normal agricultural
practice would be returned back to the fields. An alternative, non-dispersive
use of the biomass has not yet been developed for this example in Belarus.
It may hence be concluded that, except for 90Sr, annual removal ofcontaminants with plant biomass is generally too low to allow phytoextraction
to be efficient without soil additives that increase bioavailability. The highremoval rates in agricultural crops for strontium suggest that phytoextraction
could be explored with benefit for this element.
6.10.4. Caesium removal
Given its similarity to potassium, the soil potassium status will affect 137Cs
availability [141, 426–430]. Generally, the higher the soil potassium, the lower
the TF. Extremely low soil fertility with regard to potassium may increase 137Cs
TFs tenfold to 100-fold [427, 431, 432], but will also decrease plant growth. A
decrease in pH [136, 426] and decreased ammonium levels [141, 398, 433–436]
generally increase caesium soil to plant transfer, but the effects are generally
limited (a factor of 2).

69

TABLE 8.
STRONTIUM-90 ACCUMULATION BY CLOVER ON
PODZOLUVISOL LOAMY SAND SOIL IN BELARUS (DEPOSITION:
37 kBq/m2) [425]Fertilizer
treatment

Crop yield(green mass)(t/ha)

90
90

Sr activity(Bq/kg)Sr accumulationin yield (kBq/ha)

Sr accumulation(% of total soil
concentration)

90

For soil pH (KCl) = 4.9
P60

36

243

8 809

2.4

P60K6046

238

10 948

3.0

P60K12054

223

12 098

3.3

P60K18065

207

13 455

3.6

P60

40

198

7 871

2.1

P60K60

49

188

9 165

2.5

P60K12057

178

10 191

2.8

P60K180

72

160

11 4403.1

For soil pH (KCl) = 5.9

For soil pH (KCl) = 6.8
P60

46

169

7 732

2.1

P60K60

55

153

8 339

2.3

P60K120

64

151

9 589

2.6

P60K180

75

123

9 194

2.5

The effect of ammonium addition on the phytoextraction potential ofryegrass and Brassica grown on caesium contaminated soil has been tested[398]. Ammonium addition increased the dry weight yield by 20% and the TF
by 80%, resulting in a TF of 0.8 g/g. With a realistic yield of 20 t/ha under field
conditions, this would result in an annual reduction of 3.3% (decay included).
This would imply in turn that 50 years of continued phytoextraction would beneeded to reach a reduction of the soil contamination level by a factor of 5 (cf.
Table 6).
Amarantus species have TFs as high as 3.2 g/g [437, 438]. With a yieldpotential estimated at around 30 t/ha/a (based on two harvests per year) and atarget fourfold reduction in soil activity, the phytoextraction process would
require 16 years to complete.

70

In a normal agricultural land use system the annual 137Cs removal withplant yield is rather small compared with the total amount of contamination
derived 137Cs present in the soil [439]. It is known that the highest caesium
uptake typically occurs in perennial grasses. As found recently in several field
experiments in Belarus, 137Cs removal rates for perennial grasses with anannual dry matter yield of 2–5 t/ha are less than 0.1% [425]. Phytoextraction of
caesium in normal agricultural practice therefore appears not to be a very
efficient process.
6.10.5. Phytoextraction project in Belarus
The phytoextraction effect of rape (Brassica sp.) is significant. Rape has ahigh ability to accumulate 90Sr [126, 239]. In BRISSA field experiments theannual accumulation of 90Sr in pods and straw reached approximately 3% of
the 90Sr content in the soil (Table 9). Radionuclides incorporated in straw
ploughed in just after harvesting will be unavailable for one to two subsequent
growing seasons, until the final mineralization of the straw. The degree of 90Sr
immobilization by straw is comparable in size to the reduction of soil contamination due to radioactive decay. The phytoremediation effect of growing rape
may be increased by removing straw from the field and disposing of it safely.
However, the disposal option is likely to be rather expensive and will deprive
the soil of the necessary raw material for humus formation. Thus while
phytoremediation with rape appears feasible in principle, it might be moresustainable to operate the scheme as a means for enhanced attenuation.
Available data indicate a significant interspecies variability in the transferof radionuclides from soil to plants. However, hard experimental data for theevaluation of phytoextraction potential and for the development of an
appropriate crop rotation scheme are rather scarce. Experimental data from
Belarus show differences in the accumulation of 137Cs for 32 varieties of springrape between years of up to 1.8–2.7 times, and for 90Sr of up to 1.8–4.0 times. It
should be noted that these differences are radionuclide specific, meaning thatone variety that accumulates less 137Cs does not necessarily accumulate less 90Sr
[126]. The experimental results from Belarus allow the identification ofvarieties that have the desired uptake properties: more uptake for phytoextraction purposes or less uptake for minimizing the radionuclide content in
the food pathway [134, 239].

71

TABLE 9. STRONTIUM-90 ACCUMULATION BY VARIETIES OFSPRING RAPE RELATED TO PODZOLUVISOL LOAMY SAND SOIL
WITH A DEPOSITION OF 37 kBq/m2 (1997–1998) [126]90

Sr activity(Bq/kg)90

Sr uptake(kBq/ha)90

Sr uptake(% of total soil content)

Seed
yield
(t/ha)

Seeds

Straw

Seeds

Straw

Seeds

Straw

Total

Hanna
(standard)

1.9

265

663

514

5 141

0.14

1.39

1.53

Yavor

1.9

240

648

451

4 873

0.12

1.32

1.44

Likosmos

2.0

264

686

541

5 628

0.15

1.52

1.67

Lirovel

1.8

275

688

501

5 005

0.14

1.35

1.49

Licoll

2.2

310

868

682

7 638

0.18

2.06

2.25

PF 7118/93

2.2

314

879

675

7 561

0.18

2.04

2.23

PF 7045/91

2.1

319

479

670

4 019

0.18

1.09

1.27

PF 7056/92

1.9

322

902

570

6 383

0.15

1.73

1.88

Iris

1.9

338

744

629

5 532

0.17

1.50

1.67

Orakel

2.1

344

826

726

6 968

0.20

1.88

2.08

PF 5045/88

2.2

345

759

742

6 527

0.20

1.76

1.96

PF 7369/94

2.3

358

967

816

8 815

0.22

2.38

2.60

Lizonne

1.7

395

1 027

675

7 025

0.18

1.90

2.08

PF 7410/94

1.9

407

1 140

765

8 570

0.21

2.32

2.52

Liazon

2.0

436

1 221

855

9 571

0.23

2.59

2.82

PF 7041/91

1.8

477

1 336

844

9 456

0.23

2.56

2.78

PF 7008/91

2.4

478

1 338

1 166

13 063

0.32

3.53

3.85

Variety

6.11. RHIZOFILTRATION
Rhizofiltration is the use of plants to sequester compounds from aqueous
solutions through adsorption on the roots or assimilation through the roots andeventual translocation to the aerial biomass (phytoextraction). Rhizofiltration
is being investigated for the removal of radionuclides from aqueous waste
streams, including groundwater and wastewater [440].
Rhizofiltration is particularly effective in applications with low concentrations and large volumes of water. Plants that are efficient at translocating72

metals to the shoots should not be used for rhizofiltration, since additional
contaminated plant residue is produced [441].
The removal of a radionuclide from an aqueous waste stream is governedby the plant dry weight production and the concentration factor, CF (ratio of
Bq/g plant to Bq/mL water or soil solution). Since adsorption in (waste)water
per volume is lower than in soil, the CF is higher than the TF. This becomes
clear when considering the relationship between the TF and the CF, which is
TF = CF/KD, in which KD is the solid–liquid distribution coefficient of a radionuclide (e.g. dm3/kg) (i.e. the ratio of radionuclide activity concentration in thesolid phase to that in the soil solution). Since the value of KD for most radionuclides is generally substantially higher than 1 [27, 442, 443], it is clear that the
CF exceeds the TF by the same factor and that rhizofiltration is generally more
effective than soil phytoextraction.
A plant suitable for rhizofiltration applications can remove toxic metals
from solution over an extended period of time with its rapid growth rootsystem. Various plant species have been found to effectively remove toxic
elements such as arsenic, copper, cadmium, chromium, nickel, lead and zincfrom aqueous solutions [444, 445].
Pilot scale research on rhizofiltration has found that the roots ofsunflowers (Helianthus annuus L.) reduced levels of lead, copper, zinc, nickel,
strontium, cadmium, U(VI), manganese and Cr(VI) to concentrations near toor below regulated discharge limits within 24 h [441, 446]. Beans and mustard
were less effective than sunflowers in uranium removal [447]. Virtually all
uranium was concentrated in the roots, and almost none in the shoots. Removal
was higher (by a factor of 2) at pH5 than at pH7.
Uranium is clearly removed much faster from contaminated pond water
than caesium and strontium (Fig. 23) [447]. Sunflowers showed higher caesium
and strontium removal rates than timothy, meadow foxtail, Indian mustard andpeas.
However, rhizofiltration has its limits [447]. In an experiment with rather
highly contaminated wastewater (1 mg/L U) and high flow rates (1.05 L/min),
95% of the uranium was removed by 6 week old sunflowers grown for 2 weeks
in the wastewater, resulting in effluent concentrations of 40–70 µg/L, which areabove the 20 µg/L drinking water limit.

73

(a)

U (µg/L)

60

40

20

0
0

20

40

60

80

Time (h)SF 187SF Mammoth

Brassica juncea

Beans

Control

600

Metal concentration (µg/L)(b)

Cs (µg/L)
Sr (µg/L)

500

U (µg/L)
400
300
200
100
0
0

20

40

60

80

100

Time (h)FIG. 23. Removal of uranium by different sunflower cultivars (a) and removal of
caesium, strontium and uranium by sunflowers (b) [447].

74

7. SUMMARY AND CONCLUSIONSDispersed low level contamination poses a particular challenge to thosecharged with its remediation. Many techniques are not efficient below certain
concentration thresholds or entail more severe impacts on certain environmental compartments than the contamination itself. In such cases justification
for remediation may not be given on radiation protection grounds, but
remediation may still be demanded by the public.
This report examines a variety of technological options for dealing withdispersed low level contamination. The objective of any technology used in a
remediation project is either to remove or reduce the source term or to block
exposure pathways. This can be achieved in a variety of ways and needs to be
tailored to the contaminants and pathways of interest. The approaches arebroadly grouped into the three categories of:
(a)
(b)
(c)

Non-intervention;
Containment;
Removal.

While this conceptual grouping is useful for understanding the objectives
behind certain technologies, it may be noted that various technologies could begrouped into more than one category. The functionality of a certain technology
may change over time or may combine, for example, containment withremoval.
The decision of whether to remediate actively would be based on
predictive modelling and a risk assessment. If there are no acute exposure
pathways, the decision not to intervene may be justified. However, the effects
of natural attenuation would be monitored to verify that the contamination
behaves as predicted.
Reliance on natural processes may need to be complemented by a change
in the use of the contaminated land in order to minimize exposures and theuptake of radionuclides. Such change in land use may extend to (temporarily)
dedicating agricultural land to forestry or may simply involve a change of crops.
In coping with the widespread contamination following the Chernobyl
accident, such agricultural countermeasures have found wide application forretaining some use of the land.
There are various ways by which enhanced attenuation can be attempted.
Typically such measures intend to change the controlling geochemicalparameters, such as pH or redox potential, to values that would disfavour

75migration. Enhancement of sorption binding capacities and intensities by soil
additives has also been attempted.
Such measures gradually lead to more invasive measures to contain acontamination. Outright containment by impermeable or low permeability
barriers and liners may encounter logistic problems for dispersed contamination. Apart from the fact that this would involve major civil engineering
work, the effectiveness may be difficult to maintain.
Permeable reactive barriers seem to offer a viable alternative for a wide
range of contaminants. Such barriers are intended to fix in situ a contaminantpassing by with the groundwater flow. The fixation mechanisms typically aresorption or precipitation, or a combination of both. Frequently, the reactants inthe barrier are also meant to induce a change in the redox of the radionuclide(or heavy metal) concerned, leading to a less mobile species. Various
construction methods are employed, including trenching and injection curtains.
A rather more novel application are biowalls that utilize microbial processes toeffect fixation.
It may also be possible to immobilize a contaminant in situ, rather than
waiting for it to migrate into a reactive barrier. Technologies involve, forexample, either the injection of grouts to bind contaminants and to reduce soil
permeability, or the injection of reactants to precipitate the radionuclides. In
the latter case, the reactant may be compressed air or another oxidant tooxidize the iron in the groundwater to ferric oxyhydroxides, which act as asoprtion substrate. Compared with reactive barriers, however, the amount ofcivil engineering work is likely to be higher.
In certain environments the dispersal of surface contamination in theform of contaminated dust may be of concern. Organic polymers and otherbinders have been successfully used to suppress dust generation. Reducing
erodibility may also be an important first step in establishing a stable
vegetation cover. In addition to mechanically stabilizing the topsoil, phytostabilization can establish a well determined cycling of radionuclides in a limited
ecosystem, thus preventing their further dispersal.
Plant systems and more specifically aqueous plant systems are actively
used to sequester radionuclides from surface waters, including mine drainage
and drainage from waste disposal facilities. While such (constructed) wetlands
have a certain attraction in requiring relatively little attention once established,
the longevity of the solution very much depends on the stability of theecosystem created. When, for example, the wetland dries out, there is a risk of
the radionuclides trapped in the root system or the shoots becoming remobilized. In climates with severe winters, such constructed wetlands may only
function during the summer vegetation period.

76The baseline techniques for contamination removal from soils andgroundwater are excavation and pump and treat, respectively. While
excavation removes the contamination, if carried out judiciously, it creates
equivalent volumes of material that must be treated and disposed of. In thecase of dispersed contamination this may potentially lead to large volumes ofwaste and hence may not be feasible. The excavated material also may have tobe replaced with other, uncontaminated material, for which no resource may
be available.
Pump and treat methods are based on the assumption that the source canbe completely removed by physical displacement. Experience and predictivemodelling of the controlling processes have, however, shown that almost
always a residual contamination will remain. This residual contamination will
manifest itself as a long concentration ‘tail’ under a steady pumping regime andmay result in rising concentrations when pumping is stopped or interrupted.
For this reason, nowadays pump and treat is considered more a (dynamic)
containment technique than a real removal technique. This probably applies
even if enhanced recovery techniques are applied, such as injection of lixiviant
into an aquifer. Likely recovery rates can be estimated taking into account theexperience of the oil industry and the mining industry employing in situ
leaching [378]. More than half of the original concentrations of the target
material (oil, metal value) may remain underground. Leaching techniques havealso been applied to the vadose zone, but little information exists on theirefficiency for the removal of non-biodegradable materials.
If applied, ex situ extraction techniques will be able to overcome hydrodynamic limitations, by agitating the excavated material as a slurry, etc.
Arrangements such as heap leaching have the advantage that the soil or rock
structure can be broken down and thus the limitations due to inhomogeneities
can be overcome. This may not hold, however, at the microscopic level. Fine
grained material may need to be disintegrated considerably to achievereasonable recovery rates.
Heap leaching, like various other leaching methods, may involve (micro-)
biological components. Bioleaching can be actively promoted by stimulating
the growth of specialized microorganisms. These microorganisms influence
their geochemical environment and may involve the radionuclides in their
metabolic activity, which leads to a mobilization of the radionuclides. The main
advantage of bioleaching is seen in obviating the use of large quantities oflixiviants, such as strong acids, and the ensuing need to dispose of theselixiviants once the remediation is completed.
The efficiency of any ex situ removal process can be significantly
enhanced by appropriate characterization and preceding segregation ofcontaminated materials. Inorganic and organic contaminations tend to be77

preferentially associated with the fine fraction of soils, and hence a grain size
separation will greatly reduce the amount of material that requires treatment.
It is important to note that any extractive technique applied to a topsoil islikely to impair seriously its functionality and fertility, resulting in a sterilematerial.
Over the past few decades many processes for removing radionuclides
and other metals from solution have been developed. In addition to the
traditional precipitation, sorption and ion exchange methods, various biotechnology methods, such as biosorption, have been developed. Although data onlarge scale application in a remediation context are still lacking, the experience
gained from metallurgical applications is promising.
Higher plants can also be used to extract metals, including radionuclides,from soil. The uptake is very plant and radionuclide specific. After the
Chernobyl accident, plant uptake by various agricultural species was
extensively studied, with a view both to identifying minimal uptake conditions
and to maximizing uptake in a phytoextraction context. The data showed thatthe rate of sequestration would be 1–2% of the total soil reservoir at best per
crop cycle, resulting in proportionately long treatment times. Addition ofcomplexing agents to the soil can help mobilize radionuclides such as uranium
and facilitate plant uptake, resulting in sequestration rates of up to 5%. The
potentially negative effect of such additives on plant growth and biomass yield,
however, has to be carefully evaluated. The effective depth of phytoextraction
is limited by the penetration depth of the roots, which is typically the tilling
depth in agricultural soils. Adequate uses or disposal routes for the harvested
contaminated biomass need to be found.
Another technique involving higher plants to sequester radionuclides
from solution is rhizofiltration, in which the geochemical environment createdby plant roots is utilized to remove radionuclides from solution. A variety ofphysicochemical processes may actually be at work. Radionuclides that are lessmobile under reducing conditions, such as uranium, and that sorb more readily,
are more accessible to fixation by rhizofiltration than radionuclides that belong
to the alkaline or earth alkaline groups (caesium and strontium).
This report shows that a wide variety of remediation techniques havebeen developed over the years. Dealing with dispersed radionuclides in the
topsoil, however, remains difficult. There are serious trade-offs to be madebetween the degree of invasiveness and the efficiency of a technique. The sheer
volume of material to be treated often precludes the use of certain techniques
known to be efficient. Where direct exposure is not a concern, techniques andstrategies that prevent further dispersion, and hence limit or interrupt
pathways by which the radionuclides would enter the food pathway, seem to be
the most appropriate.

78

Similarly, removal of radionuclides from groundwater tends to be rather
inefficient. Containment to prevent further dispersal, for example by reactive
barriers, appears to be a better choice.
In all cases where direct exposure is not a concern, monitored naturalattenuation should be the baseline against which active remediation measures
should be evaluated.

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GLOSSARY
AMD. Acid mine drainage.
ARD. Acid rock drainage.
ATP. Adenosine triphosphate.
CF. Concentration factor for a radionuclide between different environmental
compartments, for example soil solution and plants (ratio of Bq/g plant to
Bq/mL water or soil solution).
electrowinning. Concentration of metals from a pregnant solution using
electrolysis techniques.
NADH. Nicotinamide adenine dinucleotide.
pregnant. Said of metal bearing leach solutions after contact with the ore.
TF. Transfer factor for a radionuclide between different environmental
compartments, for example soil and plants. Unit depends on the original(activity) concentration units for the respective compartments.
SRC. Short rotation coppice. Woodland management scheme whereby rapidly
growing tree species are cut every few years to harvest the biomass.

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CONTRIBUTORS TO DRAFTING AND REVIEW
Blowes, D.

University of Waterloo, Canada

Bogdevitch, I.

Belorussian Research Institute for Soil Sciences andAgrochemistry, Belarus

Falck, W.E.

International Atomic Energy Agency

Jova Sed, L.

International Atomic Energy Agency

Read, D.

University of Reading, United Kingdom

Regens, J.L.

Oklahoma University, United States of America

Sand, W.

Universität Hamburg, Germany

Tsezos, M.

National Technical University of Athens, Greece

Vandenhove, H.

SCK•CEN, Belgium

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