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4.5.8 In-sito technologies (SRT)

Contents In-situ treatment for contaminant destruction and removal In-situ biological remediation Phyto-remediation Phyto-extraction treatment Uranium removal Strontium removal Caesium removal Example: Phyto-extraction project in Belarus Rhizo-filtration treatment Non-biological in-situ treatment Dynamic underground stripping and hydrous pyrolysis oxidation Soil vapour extraction Thermal methods In-situ treatment for contaminant destruction and removal

Treatment technologies are source control technologies that reduce the toxicity and/or volume of the waste by destroying or removing polluting constituents. Treatment technologies are capable of permanently reducing the overall risk posed by wastes.

In-situ treatment technologies allow soil or groundwater to be treated without being excavated and transported, resulting in potentially significant cost savings. However, in-situ treatment generally requires longer time periods than ex-situ treatment, and there is less certainty about uniformity of treatment because of the variability in soil and aquifer characteristics and because the effectiveness of the process is more difficult to verify. The major categories of in-situ treatment processes are biological, physical/chemical, and thermal treatment. In-situ treatment technologies are generally not applicable to bulk waste.

A single technique may not be sufficient for the remediation of a situation with mixed contamination. In the following a range of techniques that specifically address organic contaminants are described that would be complementary to other techniques addressing, for instance, heavy metals and radionuclides:

  1. Biological remediation whereas a separate section is addressed to phyto-remediation.
  2. Dynamic underground stripping and hydrous pyrolysis oxidation;
  3. Soil vapour extraction;
  4. Thermal techniques, such as: electrical resistance heating, microwave heating or thermal conductance.

The last 3 groups of non-biological techniques are characterised in detail in Section

h4 In-situ biological remediation

In general, bio-remediation technologies employ engineered systems to heighten the effects of naturally occurring degradation mechanisms [IAEA-2006b]. Bio-remediation techniques are destruction or transformation techniques directed towards stimulating micro-organisms to grow by using the contaminants as a food and energy source through creating a favourable environment for the micro-organisms. In general, this means providing some combination of oxygen, nutrients and in some cases moisture, and controlling the temperature and pH. Sometimes, micro-organisms adapted for degradation of the specific contaminants are applied to enhance the process. There is a conceptual similarity to techniques used in the context of enhanced natural attenuation, see Section

Bio-degradation methods are likely to gain ground, as disposal related legislation increasingly tends to discourage or prohibit landfilling of bio-degradable materials. However, the application of bio-remediation techniques, although often cost efficient, may be hampered by licensing procedures [IAEA-2006b].

The rate at which micro-organisms degrade contaminants is influenced by the following parameters:

  • Specific contaminants present;
  • Oxygen supply: in aerobic conditions mechanical tilling, venting or sparging are used; anaerobic conditions may be used to degrade highly chlorinated contaminants;
  • Moisture: levels in the range of 20 – 80 % generally allow suitable bio-degradation in soils;
  • Nutrient supply: if nutrients are not available in sufficient amounts, microbial activity will stop; nitrogen and phosphorus are the nutrients most likely to be deficient in the contaminated environment and are thus usually added to the bio-remediation system in a useable form (e.g., as ammonium for nitrogen and as phosphate for phosphorus);
  • pH affects the solubility, and consequently the availability, of many constituents of soil, which can affect the biological activity; many metals that are potentially toxic to micro-organisms are insoluble at elevated pH levels; therefore, elevation of the pH of the treatment system can reduce the risk of poisoning the micro-organisms;
  • Temperature: the bio-degradation rate will slow down with decreasing temperature;
  • Availability of the contaminant to the micro-organism (clays can adsorb contaminants, making them unavailable to micro-organisms);
  • Concentrations of the contaminants (high concentrations may be toxic to micro-organisms);
  • Presence of substances toxic to micro-organisms, e.g., mercury or inhibitors to the metabolism of the contaminant.

These parameters are discussed briefly in the following and also pertain to ex-situ methods.
A wide variety of process designs and technical layouts have been developed. These may be based on groundwater recirculation (Figure 4.25) or direct injection (Figure 4.26) [IAEA-2006b]. The features of the above mentioned techniques are shown in Table 4.6, Table 4.7 and Table 4.8 within Section 4.5.

Figure 4.25 Stimulation of in-situ bio-remediation by groundwater recirculation
Figure 4.25 Stimulation of in-situ bio-remediation by groundwater recirculation


Figure 4.26 Bio-remediation by direct injection of nutrients
Figure 4.26 Bio-remediation by direct injection of nutrients

Natural micro-biological systems are very complex, difficult to understand in their interactions, and, unlike many engineered systems, difficult to control. In this sense, bio-remediation is not foolproof and it cannot be guaranteed to be successful even in instances where due care was taken in its design and application.

Micro-biologically specific reasons for the poor performance of in-situ bio-remediation systems include [IAEA-2006b]:

  1. There is uncertainty with regard to the effect of hydrocarbon availability on the effectiveness of bio-degradation. Can bacteria degrade hydrocarbons adsorbed on surfaces or degrade hydrocarbons with low levels of solubility? Or must the hydrocarbon be solubilized before it can be bio-degraded?
  2. Although petroleum hydrocarbons are amenable to aerobic bio-degradation, for it to occur the indigenous bacteria must have the appropriate genetic information. This genetic information is precise. The presence of a specific hydrocarbon will stimulate the synthesis of an oxy-genase enzyme that is expressly configured to react with that stimulating hydrocarbon. For remediation, the question is whether the indigenous microbes possess the genetic information required for appropriate enzyme production and whether the contaminant stimulates the production of those enzymes.
  3. General microbial stimulation has the potential to produce a large amount of biomass that may not take part in the bio-degradation process and actually be harmful through bio-fouling and plugging of injection wells, galleries or surrounding formations. There is potential to lose critical subsurface mass transport capabilities.
  4. There are practical limits to the degree of clean-up obtainable using bio-remediation. Hydrocarbons at the low ppm level may not be capable of supporting significant levels of microbial activity even under stimulation. Sites with relatively high levels of hydrocarbon impact may actually be better candidates for bio-remediation than those on which the impact is small at levels slightly above regulatory action levels.

It should be noted that many of these factors are better controllable under ex-situ conditions described in Section 4.5.9. Phyto-remediation

In-situ bio-remediation may also employ higher plants and is then commonly known under the title phyto-remediation. Here the contaminants are either taken up into the shoots or the roots, or the complex bio-geochemical processes in the root zone either destroy or immobilize the contaminants.

Studies on the efficiency of bio-degradation in the presence of radionuclides and heavy metals are important, since metabolic pathways can be inhibited in their presence. Some fungi have been shown to be tolerant to high metal concentrations. Laboratory research also indicates that fungi that are resistant to metals in symbiotic association with plant roots might positively influence phyto-remediation [IAEA-2006b]. An overview of typical phyto-remediation techniques and their applicability to individual type of media and various target contaminants as well as their respective state of development is shown in Table 4.6, Table 4.7 and Table 4.8 within Section 4.5.

Most relevant research has focused on individual contaminants or on certain classes of contaminant and not on mixtures of different types of contaminant.

Details on the phyto-extraction method and its application for uranium, strontium and caesium removal from soil as well as results from a phyto-extraction project in Belarus are given in Section

The Rhizo-filtration method as another example of a phyto-remediation technique suitable for groundwater remediation is described in Section Phyto-extraction treatment

The use of plants to remove contaminants from the environment and concentrate them in above ground plant tissue is known as phyto-extraction. Phyto-extraction requires that the target metal (radionuclide) be available to the plant root, absorbed by the root and trans-located from the root to the shoot; biomass production should be substantial. The metal (radionuclide) is removed from the site by harvesting the biomass, after which it is processed either to recover the metal or further concentrate the metal (by a thermal, microbial or chemical treatment) to facilitate disposal [IAEA-2004b].

Research and development efforts have focused on two areas:

  1. Remediation of contaminants such as lead, arsenic, chromium, mercury and radionuclides; and
  2. Mining, or recovery, of inorganic compounds, mainly nickel and copper, having intrinsic economic value [IAEA-2004b].

Successful implementation of phyto-extraction depends on [IAEA-2004b]:

  • The bio-availability of the contaminant in the environmental matrix;
  • Root uptake;
  • Internal translocation of the plant;
  • Plant tolerance.

Plant productivity (i.e., the amount of dry matter that is harvestable each season) and the accumulation factor (the ratio of metal in plant tissue to that in the soil) are important design parameters. This is clearly exemplified by the following set of equations and tables. The percentage yearly reduction in soil activity can be calculated as:

Annual removal (%) = ( (TF x yield) / Wsoil ) x 100 …………………………………… 1


TF = the transfer factor or bio-accumulation factor (TF = Cplant/Csoil)
Cplant = the concentration of the radio-contaminant in the plant (Bq/g)
Csoil = the concentration of the contaminant in the soil ( )
Wsoil = the weight of the contaminated soil layer (kg/ha)

As is evident from this equation, the annual removal percentage increases with yield and transfer factor. However, transfer factor and yield values are not independent: a high yield is often associated with lower transfer factors because of growth dilution effects.
Phyto-extraction typically requires several years of operation, and the future trend in radionuclide concentration in the soil can be calculated from:

Csoil,t = Csoil,t=0 exp{ – ( (TF x yield) / Wsoil + 0.69 / t1/2 ) x t} ………………… 2

The second term in the exponent of this equation 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 the phyto-extraction potential. For others, for example 137Cs and 90Sr, with half-lives of 30 years, the phyto-extraction potential will be affected; that is, a yearly loss of 2.33 % in activity occurs merely through radioactive decay. The equation (2) assumes a constant bio-availability of the contaminant (i.e., a constant transfer factor (TF)).
Table 4.10 shows, for a calculated example, the percentage annual removal per hectare for different crop yields and transfer factors, based on a 10 cm deep soil layer that has a mass of 1500 t for a soil density of 1.5 kg/dm3. It should be borne in mind 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 4.12.

TF [g/g] Annual reduction
due to phytoextraction [%]
Annual reduction
due to phytoextraction and decay [%]
Yield (t/ha) Yield (t/ha)
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

Table 4.10 Percentage yearly reduction of soil contamination due to phyto-extraction and radioactive decay ( Note: t1/2: 30a; soil depth: 10 cm; soil density: 1.5 kg.dm3).

Nuclide Total range
(Bq/g plant to Bq/g soil)
Comment on conditions for upper limit
Cs 0.00025 – 7.5 Brassica, organic soil
Sr 00051 – 22 Green vegetables, sandy soil
U 0.000006 -21.13 Tubers, sandy soil
Ra 0.00029 – 0.21 Grass, sandy soil

Table 4.11 Ranges for transfer factors (ratio) based on data from references

Yields of more than 20 t/ha and transfer factors higher than 0.1 (Table 4.10) may be regarded as upper limits, except for strontium. This would result in an annual reduction percentage of 0.1 % (decay excluded). When the transfer factor equals 1, the annual reduction is about 1 %. Table 4.11 gives some ranges for transfer factors for the natural radionuclides uranium and radium, predominant contaminants in the natural occurring radioactive materials (NORM) industries, and the long lived fission products 137Cs and 90Sr.

By rearranging equation (2), the number of years needed to attain the required reduction factor as a function of annual removal percentage can be calculated. Table 4.12 presents the number of years required to attain a reduction of the contaminant concentration up to a factor of 100, given an annual extraction 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 that measures would need to be taken to increase the annual removal efficiency through crop selection, or to increase the bio-availability by applying soil additives and through technical measures (e.g., decreasing the tilled soil depth).

factor ( )
Activity 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

Table 4.12 Calculated number of years required to decontaminate a soil for a required desired) reduction factor and a given annual removal percentage (Note:Soil depth: 10 cm; soil density: 1.5 kg/dm3)

In most cases one has limited control over the depth of the contamination, although it may be feasible and advantageous to excavate and pile the soil to the desired soil depth for phyto-remediation purposes. One possibility is to excavate the soil and spread it on geo-membranes, which impedes roots from penetrating to deeper layers. These membranes will also limit contaminant dispersal to the underlying clean soil, but a substantial area may be needed for treatment. Decreasing the tilled soil depth increases the removal percentage according to equation (1), and may intensify root-soil contact, and may result in an increased transfer factor.

The other factors influencing radionuclide bio-availability, such as crop selection and measures to increase the bio-availability of the radionuclide of concern, 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, or the addition of agents that increase the availability of metals.

Apart from the application of soil additives to increase export with the plant biomass, plant selection may also be important for improving the phyto-extraction potential. As already mentioned, there is a significant interspecies variability in transfer factors (Table 4.15). Since the values are seldom obtained for similar soil and growth conditions, the effect of plant species on the transfer factors cannot be unambiguously derived. Observed differences between plant varieties or cultivars have been up to a factor of 2 [IAEA-2004b].

Improved genotypes with optimized metal uptake, translocation and tolerance, and improved biomass yield, may also be an approach to improved phyto-extraction. Plant breeding and genetic engineering may open further alleys to develop hyper-accumulating plants, but actual research and technology development is mostly limited to heavy metals [IAEA-2004b].

Although positive effects have been obtained following applications of soil amendments that increase element bio-availability, the effect of continuous treatment on soil quality, plant growth and bio-accumulation is not clear. There also remains the question of long term effectiveness: will transfer factors remain constant or will they decrease as radionuclide concentrations decrease.

Effective extraction of radionuclides and heavy metals by hyper-accumulators is limited to shallow soil depths of up to 30 cm. If a contamination is found 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 the contaminated zone would be small and hence also the phyto-extraction potential.

There are concerns that contaminated leaf litter and associated toxic residues 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 phyto-extraction scheme [IAEA-2004b].

Little is known about the economics of phyto-extraction, 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 for annual 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 to 20 kg of biomass (~ 2 – 4 kg of ash) annually. Uranium removal

Free UO22+ is the uranium species most readily taken up and translocated by 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 phyto-extraction. The uranyl cat-ion also binds to the soil solids and organic matter, reducing the extent of plant uptake. Therefore, in addition to acidification, soil amendments that increase the availability of uranium by complexation may also be required. In testing the role of acidification and chelating agents on the solubilization of uranium it was found that, of the organic acids and chelating agents tested, citric acid was the most effective for increasing uranium in the soil solution. Following citric acid treatment (20 mmol/kg) the uranium accumulation in Indian mustard (Brassica juncea) was increased 1000-fold and in beet (Beta vulgaris) tenfold [IAEA-2004b].

Similar results were obtained when testing the potential for phyto-extraction of uranium from a low level contaminated sandy soil using rye grass (Lolium perenne cv. Melvina), Indian mustard (Brassica juncea cv. Vitasso) and redroot pigweed (Amarathus retroflexus) [IAEA-2004b]. 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 re-growth (in the case 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 4.12). Strontium removal

Table 4.13 shows the annual crop removal of 137Cs and 90Sr. It is clear from this 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 all less than 1 %, and the highest removal is found for grassland. The high sorption of 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 the availability of 90Sr is typically tenfold above that of caesium. The transfer factors of 90Sr in green vegetables and Brassica plants are typically around unity and the upper levels are around 10. Phyto-extraction of 90Sr has not yet been investigated at the field scale. The high removal rates in agricultural crops (Table 4.13) suggest that phyto-extraction may be worth while to explore [IAEA-2004b].

The highest transfers of 90Sr are typical for leguminous perennial grasses (Trifolium family) and Brassica plants. Field experiments in Belarus were carried out at the Belarussian Research Institute for Soil Science and Agrochemistry (BRISSA) on light-textured soil contaminated with 90Sr (Table 4.14). 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 were in the range 2.5 – 3.6 % of the total radionuclide reservoir in the soil. A change of soil 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 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, the greater 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 [IAEA-2004b].

(dry matter)
Ceasium TF
Ceasium crop removal
(% of total in soil)
Strontium TF
Strontium crop removal
(% of total in soil)
5-7 0.0004-0.25 0.0005-0.06 0.02-0.94 0.0037-0.22
6-10 0.003-0.89 0.0006-0.3 0.03-1.4 0.006-0.5
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

Table 4.13 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)

Crop yield
(green mass)
90Sr accumulation
in yield (kBq/ha)
90Sr accumulation
(% of total soil

For soil pH (KCl) = 4.9
P60 36 243 8809 2.4
P60K60 46 238 10948 3.0
P60K120 54 223 12098 3.3
P60K180 65 207 13455 3.6

For soil pH (KCl) = 5.9
P60 40 198 7871 2.1
P60K60 49 188 9165 2.5
P60K120 57 178 10191 2.8
P60K180 72 160 11440 3.1

For soil pH (KCl) = 6.8
P60 46 169 7732 2.1
P60K60 55 153 8339 2.3
P60K120 64 151 9589 2.6
P60K180 75 123 9194 2.5

Table 4.14 90Sr accumulation by clover on podzoluvisol loamy sand soil in Belarus (deposition: 37 kBq/m2) [IAEA-2004b]

It may hence be concluded that, except for 90Sr, annual removal of contaminants with plant biomass is generally too low to allow phyto-extraction to be efficient without soil additives that increase bio-availability. The high removal rates in agricultural crops for strontium suggest that phyto-extraction could be explored with benefit for this element. Caesium removal

Given its similarity to potassium, the soil potassium status will affect 137Cs availability. Generally, the higher the soil potassium, the lower the transfer factor. Extremely low soil fertility with regard to potassium may increase 137Cs transfer factors tenfold to 100-fold, but will also decrease plant growth. A decrease in pH and decreased ammonium levels generally increase caesium soil to plant transfer, but the effects are generally limited (a factor of 2) [IAEA-2004b].

The effect of ammonium addition on the phyto-extraction potential of ryegrass and Brassica grown on caesium contaminated soil has been tested. Ammonium addition increased the dry weight yield by 20 % and the transfer factor by 80 %, resulting in a transfer factor 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 phyto-extraction would be needed to reach a reduction of the soil contamination level by a factor of 5 (Table 4.12) [IAEA-2004b].

Amarantus species have transfer factors as high as 3.2 g/g. With a yield potential estimated at around 30 t/ha/a (based on two harvests per year) and a target fourfold reduction in soil activity, the phyto-extraction process would require 16 years to complete [IAEA-2004b].

In a normal agricultural land use system the annual 137Cs removal with plant yield is rather small compared with the total amount of contamination derived 137Cs present in the soil. 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 an annual dry matter yield of 2 – 5 t/ha are less than 0.1 %. Phyto-extraction of caesium in normal agricultural practice therefore appears not to be a very efficient process [IAEA-2004b]. Example: Phyto-extraction project in Belarus

The phyto-extraction effect of rape (Brassica sp.) is significant. Rape has a high ability to accumulate 90Sr [IAEA-2004b]. In the BRISSA field experiments the annual accumulation of 90Sr in pods and straw reached approximately 3 % of the 90Sr content in the soil (Table 4.15). 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.

Variety Seed
90Sr activiteit
90Sr uptake
90Sr uptake
(% of total soil content)
Seeds Straw Seeds Straw Seeds Straw Total
1.9 265 663 514 5141 0.14 1.39 1.53
Yavor 1.9 240 648 451 4873 0.12 1.32 1.44
Likosmos 2.0 264 686 541 5628 0.15 1.52 1.67
Lirovel 1.8 275 688 501 5005 0.14 1.35 1.49
Licoll 2.2 310 868 682 7638 0.18 2.06 2.25
PF7118/93 2.2 314 879 675 7561 0.18 2.04 2.23
PF7045/91 2.1 319 479 670 4019 0.18 1.09 1.27
PF7056/92 1.9 322 902 570 6383 0.15 1.73 1.88
Iris 1.9 338 744 629 5532 0.17 1.50 1.67
Orakal 2.1 344 826 726 6968 0.20 1.88 2.08
PF5045/88 2.2 345 759 742 6527 0.20 1.76 1.96
PF7369/94 2.3 358 967 816 8815 0.22 2.38 2.60
Lizonne 1.7 395 1027 675 7025 0.18 1.90 2.08
PF7410/94 1.9 407 1140 765 8570 0.21 2.32 2.52
Liazon 2.0 436 1221 855 9571 0.23 2.59 2.82
PF7041/91 1.8 477 1336 844 9456 0.23 2.56 2.78
PF7008/91 2.4 478 1338 1166 13063 0.32 3.53 3.85

Table 4.15 90Sr accumulation by varieties of spring rape related to podzoluvisol loamy sand soil with a deposition of 37 kBq/m2 (1997–1998) [IAEA-2004b]

The phyto-remediation 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 phyto-remediation with rape appears feasible in principle, it might be more sustainable to operate the scheme as a means for enhanced attenuation.

Available data indicate a significant interspecies variability in the transfer of radionuclides from soil to plants. However, hard experimental data for the evaluation of phyto-extraction 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 spring rape 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 that one variety that accumulates less 137Cs does not necessarily accumulate less 90Sr. The experimental results from Belarus allow the identification of varieties that have the desired uptake properties: more uptake for phyto-extraction purposes or less uptake for minimizing the radionuclide content in the food pathway [IAEA-2004b]. Rhizo-filtration treatment

Rhizo-filtration is the use of plants to sequester compounds from aqueous solutions through adsorption on the roots or assimilation through the roots and eventual translocation to the aerial biomass (phyto-extraction). Rhizo-filtration is being investigated for the removal of radionuclides from aqueous waste streams, including groundwater and wastewater [IAEA-2004b].

Rhizo-filtration is particularly effective in applications with low concentrations and large volumes of water. Plants that are efficient at trans-locating metals to the shoots should not be used for rhizo-filtration, since additional contaminated plant residue is produced [IAEA-2004b].

The removal of a radionuclide from an aqueous waste stream is governed by 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 concentration factor is higher than the transfer factor. This becomes clear when considering the relationship between the transfer factor and the concentration factor, which is:


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 the solid phase to that in the soil solution). Since the value of KD for most radionuclides is generally substantially higher than 1, it is clear that the concentration factor exceeds the transfer factor by the same factor and that rhizo-filtration is generally more effective than soil phyto-extraction [IAEA-2004b].

A plant suitable for rhizo-filtration applications can remove toxic metals from solution over an extended period of time with its rapid growth root system. Various plant species have been found to effectively remove toxic elements such as arsenic, copper, cadmium, chromium, nickel, lead and zinc from aqueous solutions [IAEA-2004b].

Pilot scale research on rhizo-filtration has found that the roots of sunflowers (Helianthus annuus L.) reduced levels of lead, copper, zinc, nickel, strontium, cadmium, U(VI), manganese and Cr(VI) to concentrations near to or below regulated discharge limits within 24 h. Beans and mustard were less effective than sunflowers in uranium removal. 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 [IAEA-2004b].

Uranium is clearly removed much faster from contaminated pond water than caesium and strontium (Figure 4.27). Sunflowers showed higher caesium and strontium removal rates than timothy, meadow foxtail, Indian mustard and peas [IAEA-2004b].

However, rhizo-filtration has its limits. 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 are above the 20 μg/l drinking water limit [IAEA-2004b].

Figure 4.27 Removal of uranium by different sunflower cultivars (a) and removal of caesium, strontium and uranium by sunflowers (b) [IAEA-2004b]
Figure 4.27 Removal of uranium by different sunflower cultivars (a) and removal of caesium, strontium and uranium by sunflowers (b) [IAEA-2004b]

. Non-biological in-situ treatment

Non-biological in-situ treatment comprises the following technologies:

  • Dynamic underground stripping and hydrous pyrolysis oxidation
  • Soil vapour extraction
  • Thermal methods (electrical resistance heating, microwave heating, thermal conductance) Dynamic underground stripping and hydrous pyrolysis oxidation

Dynamic underground stripping and hydrous pyrolysis oxidation (DUS/HPO) is a combination of technologies that can rapidly remove organic contaminants from the subsurface where other technologies may take decades or more to achieve the desired clean-up criteria. For instance, in two field-scale applications, dynamic underground stripping and hydrous pyrolysis oxidation has achieved remediation performance in less than one tenth the time of conventional pump and treat methods, both above and below the water table, and at less overall cost [IAEA-2006b]. Major elements of the technique include steam injection, air injection, vacuum extraction, electrical resistivity heating, groundwater extraction, surface treatment of vapour and groundwater, and underground imaging and monitoring.

Dynamic underground stripping is an innovative thermal remediation technology that accelerates removal of organic compounds, both dissolved phase liquids and dense non-aqueous phase liquids (DNAPL), from soil and groundwater. In dynamic underground stripping, steam is injected at the periphery of the contaminated area to volatilize and solubilize compounds bound to the soil. Centrally located vacuum extraction wells then remove this volatilized material from the subsurface. A steam front develops in the subsurface as permeable soils are heated to the boiling point of water, and volatile organic compounds are vaporized from the hot soil. The steam sweeps the permeable zones between the injection and extraction wells. Steam injection then ceases, while the vacuum extraction continues once the front reaches the extraction wells. The vapour and any groundwater pulled through the extraction wells are treated above ground. When the steam zone collapses, groundwater re-enters the treatment zone and the steam-vacuum extraction cycle is repeated.

For application in dense clays, electrical resistive heating can also be used to enhance contaminant removal. Water and contaminants in the conductive zone are vaporized and forced into the permeable zone, being swept by the steam and then subjected to vacuum extraction.

In hydrous pyrolysis oxidation, steam and air are injected into paired wells, creating a heated oxygenated zone in the subsurface. When injection is halted, the steam condenses and contaminated groundwater returns to the heated zone where it mixes with oxygen-rich condensed steam, which destroys dissolved contaminants in-situ.

An integral component of dynamic underground stripping and hydrous pyrolysis oxidation is a sophisticated imaging system known as electrical resistance tomography (ERT), which allows real time three dimensional monitoring of the subsurface. Electrical resistance tomography is based on a cross-hole tomography system that maps changes in resistivity over time. Changes in resistivity both laterally and vertically can be related to the migration of steam through various zones between the injection and extraction wells. Electrical resistance tomography is utilized to make process adjustments to optimize the performance of dynamic underground stripping and hydrous pyrolysis oxidation.

Limitations include:

  • The process requires a large amount of energy.
  • Above ground treatment systems must be sized to handle peak extraction rates and the distribution of volatile organic compounds (VOC) in the extracted vapour and liquid streams.
  • Steam adds significant amounts of water to the subsurface, and precautions must be taken to prevent mobilization of contaminants beyond the capture zone.
  • It is not applicable at depths of less than 1.5 m; to date it has been used at depths of up to 40 m.
  • Micro-organisms destroyed by steam can foul the system, and small particles pumped to the surface can clog the system.
  • Treated soils and groundwater can remain at elevated temperatures for years after clean-up, which could affect site reuse plans. Soil vapour extraction

It may be necessary to capture and remove toxic or explosive gases before or while addressing other contaminants bound to the soil or in the groundwater. Soil vapour extraction uses a vacuum to remove volatile and some semi-volatile contaminants from the soil. The vapour-soil gas mixtures will be treated and discharged according to the applicable air discharge regulations. Extraction wells are typically used at depths of 1.5 m or greater, and have been successfully applied as deep as 90 m. Groundwater pumps may be used in conjunction with soil vapour extraction to keep groundwater from rising into the vadose zone as a result of the vacuum, or to increase the depth of the unsaturated zone. This area, termed the capillary fringe is sometimes highly contaminated, as it holds non-aqueous phase liquids lighter than water and vapours that have escaped from dissolved organic compounds in the groundwater below or from dense non-aqueous phase liquids. In soils where the contamination is deep or when there is low permeability, injecting air into the soil assists in extraction. During full-scale operation, soil vapour extraction can be run intermittently (pulsed operation) once the extracted mass removal rate has reached a steady state level. Because the process involves the continuous flow of air through the soil, it often promotes bio-degradation of low volatility organic compounds that may be present.

Soil vapour extraction can also be used ex-situ on piles of excavated soil. A vacuum is applied to a network of piping in the pile to encourage volatilization of organic compounds from the excavated media. A system for handling and treating off-gases is required.

A field pilot study is necessary to establish the feasibility of the method as well as to obtain information necessary to design and configure the system.

The soil vapour extraction technique is typically applicable to volatile organic compound and/or fuel contamination. It works only on compounds that readily vaporize (i.e., that have a high Henry’s law constant). Some limitations of the soil vapour extraction technique include:

  • A high soil moisture content requires higher vacua.
  • Soils with high organic content or soils that are extremely dry have a high affinity and retention capacity for volatile organic compounds. These conditions limit its effectiveness.
  • Soils with low permeability also limit its effectiveness.
  • Applying a vacuum to the subsurface soils can raise groundwater levels. As the soils become saturated, some contaminants may dissolve into the groundwater. As a result, groundwater can show increased contamination levels, especially at the start of this process.
  • It will not remove heavy oils, metals, polychlorinated biphenyls (PCB) or dioxins.
  • Exhaust air from in-situ soil vapour extraction systems may require treatment. Off-gas treatment is usually carried out by adsorption onto granular activated carbon.
  • It is not applicable to the saturated zone (except in the form of air sparging in wells). Thermal methods
  • Electrical resistance heating uses an electric current to heat less permeable soils such as clays and fine grained sediments so that water and contaminants trapped in these low conductivity materials are vaporized and ready for vacuum extraction. An array of electrodes is placed directly into the soil matrix and an (alternating) electric current passed through the soil, the resistance loss of which then heats the soil and the contaminants, increasing the vapour pressure of the latter. The heat also dries out the soil causing it to fracture. These fractures make the soil more permeable, increasing the removal rate of contaminants by soil vapour extraction. In addition, the heating creates an in-situ source of steam to strip contaminants from the soil, inter alia reducing the viscosity of trapped liquids and eventually allowing them to be removed by soil vapour extraction. Six phase soil heating is a typical layout that uses a low frequency electric current delivered to six electrodes in a circular array to heat soils.
    The following factors may limit the applicability and effectiveness of the process:
    • It may be self-limiting, since as the clays heat up, they dry out and the current will stop flowing [IAEA-2006b].
    • Debris or other large objects buried in the media can cause focusing of the electrical field or short-circuiting.
    • The performance is very much dependent on the type of organic substance involved and its vapour pressure, as well as the temperature and heat flows that can be achieved in the process selected.
    • There is an optimum soil moisture content as the resistance increases with decreasing moisture content and the permeability in turn decreases with increasing moisture content.
    • A low permeability will hinder the flow of steam and organic vapours towards the soil vapour extraction, thus leading to a low efficiency of the process due to the high energy input to increase vacuum and temperature.
    • Soil with a highly variable permeability may result in accessibility to the contaminated regions being uneven.
    • High soil organic matter content may reduce the efficiency of the technique due to the high affinity of organic contaminants for these constituents.
    • Air emissions will need to be controlled to be below the limits of regulatory concern or permissions may need to be sought. Off-gas treatment and permits will increase project costs.
    • Residual liquids and spent activated carbon may require further treatment or disposal.
  • Microwave heating is based on the phenomenon that dipole molecules, such as those of water, can be stimulated in their vibrational movements by high frequency electromagnetic radiation. This vibrational energy is then dissipated in the form of heat. While many organic molecules are flexible enough to adjust themselves to the electromagnetic field, they still absorb photons, which may lead to the breaking of weak bonds. Such bonds can be either within the molecule or between the molecules and a surface. Thus, microwave applications will enhance recovery of organic contaminants by either volatilizing them, by reducing the viscosity due to increased ambient temperature or by detaching them from the geomatrix [IAEA-2006b]. The microwave oven principle can be applied to soils in-situ, albeit on a grander scale.
  • Thermal conductance: In-situ thermal treatment to enhance contaminant removal can also be accomplished by a technique where heat and vacuum are applied simultaneously to soil, sediments or buried wastes. Heat flows into the soil by conduction from heaters operated at approximately 800 – 1000 ºC. Vertical thermal wells are used for deep contamination and horizontal thermal wells are used for shallow contamination. Multiple wells are installed to span the areas requiring treatment. Electric heaters are installed in the wells and wired together with power tapped from utility poles or other power sources. Vapours are extracted from some of the wells to ensure the boundaries of the heated zone are under vacuum.
    Most of the contaminant destruction occurs underground near the heat source. As soil is heated, contaminants in the subsurface are volatilized or destroyed by several mechanisms, including:
    • Evaporation;
    • Boiling;
    • Oxidation;
    • Pyrolysis;
    • Steam distillation.

Volatilized contaminants not destroyed in the subsurface are recovered and treated above ground. A wide range of soil types can be treated by this process. The high temperatures applied over a period of days result in an extremely high destruction and removal efficiency even of contaminants with high boiling points such as polychlorinated biphenyls, pesticides and other heavy hydrocarbons.
Special consideration is needed when applying this process to sites with radionuclides and or toxic metals, such as mercury, as the heating process may change the oxidation state of these contaminants, which can make them more or less mobile in the environmet.