applied and environmental soil science impact of human activities

Impact of Human Activities on Soil Contamination

Guest Editors: Fernando Jose Garbuio, Jeffrey L. Howard, and Larissa Macedo dos Santos

Applied and Environmental Soil Science

Impact of Human Activities onSoil Contamination

Applied and Environmental Soil Science

Impact of Human Activities onSoil Contamination

Guest Editors: Fernando Jose Garbuio, Jeffrey L. Howard,and Larissa Macedo dos Santos

Copyright © 2012 Hindawi Publishing Corporation. All rights reserved.

This is a special issue published in “Applied and Environmental Soil Science.” All articles are open access articles distributed under theCreative Commons Attribution License, which permits unrestricted use, distribution, and reproduction in any medium, provided theoriginal work is properly cited.

Editorial Board

Lynette K. Abbott, AustraliaJoselito M. Arocena, CanadaNanthi Bolan, AustraliaRobert L. Bradley, CanadaArtemi Cerda, SpainClaudio Cocozza, ItalyHong J. Di, New ZealandOliver Dilly, GermanyMichael A. Fullen, UK

Ryusuke Hatano, JapanWilliam R. Horwath, USAD. L. Jones, UKMatthias Kaestner, GermanyHeike Knicker, SpainTakashi Kosaki, JapanYongchao Liang, ChinaTeodoro M. Miano, ItalyAmaresh K. Nayak, India

Alessandro Piccolo, ItalyNikolla Qafoku, USAPeter Shouse, USAB. Singh, AustraliaKeith Smettem, AustraliaMarco Trevisan, ItalyAntonio Violante, ItalyPaul Voroney, CanadaJianming Xu, China

Contents

Impact of Human Activities on Soil Contamination, Fernando Jose Garbuio, Jeffrey L. Howard,and Larissa Macedo dos SantosVolume 2012, Article ID 619548, 2 pages

Acidification and Nitrogen Eutrophication of Austrian Forest Soils, Robert Jandl, Stefan Smidt,Franz Mutsch, Alfred Furst, Harald Zechmeister, Heidi Bauer, and Thomas DirnbockVolume 2012, Article ID 632602, 9 pages

Spatial Distribution of PCB Dechlorinating Bacteria and Activities in Contaminated Soil,Birthe V. Kjellerup, Piuly Paul, Upal Ghosh, Harold D. May, and Kevin R. SowersVolume 2012, Article ID 584970, 11 pages

A Critical Evaluation of Single Extractions from the SMT Program to Determine Trace Element Mobilityin Sediments, Valerie CappuynsVolume 2012, Article ID 672914, 15 pages

Distribution and Fate of Military Explosives and Propellants in Soil: A Review, John PichtelVolume 2012, Article ID 617236, 33 pages

Occurrence of Vanadium in Belgian and European Alluvial Soils, V. Cappuyns and E. SlabbinckVolume 2012, Article ID 979501, 12 pages

Hindawi Publishing CorporationApplied and Environmental Soil ScienceVolume 2012, Article ID 619548, 2 pagesdoi:10.1155/2012/619548

Editorial

Impact of Human Activities on Soil Contamination

Fernando Jose Garbuio,1 Jeffrey L. Howard,2 and Larissa Macedo dos Santos3

1 Departamento de Ciencia do Solo, Instituto Federal Catarinense, Santa Rosa do Sul, SC 48202, Brazil2 Department of Geology, Wayne State University, Detroit, MI, USA3 Department of Chemistry, Federal Technological University of Parana, Pato Branco, PR, Brazil

Correspondence should be addressed to Fernando Jose Garbuio, [email protected]

Received 17 December 2012; Accepted 17 December 2012

Copyright © 2012 Fernando Jose Garbuio et al. This is an open access article distributed under the Creative Commons AttributionLicense, which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properlycited.

The impacts of human activities on soil contamination aremany and varied. The extent of human impact is now so per-vasive and profound that there is currently much discussionabout the “Anthropocene”, a new geologic era characterizedby anthropogenic disturbances of the geologic record. Manyof the problems recognized during the 1970s linger on,including the effects of acid rain and airborne depositionof soot, fly ash, and other potentially toxic particulates.Further, the scope of the problem has grown significantlywith economic growth in previously less developed nationssuch as China and India. The effects of human activities varywith land use, ranging from agricultural wastes such as farmanimal sewerage and fertilizer runoff, to commercial andindustrial wastes of every conceivable type and magnitude.Over the years, the list of toxic contaminates has also grown,so that it not only includes heavy metals, radionuclides,and organic compounds of anthropogenic origin, but phar-maceuticals, explosives, and previously unknown biologicalpathogens. The field of soil remediation has also growntremendously over the past few decades. The goal of thisspecial issue is to further explore the effects of human activ-ities on soil contamination. Topics to be examined includethe nature and extent of soil contamination, state of theart methodologies for studying soil and related groundwatercontamination, and innovative techniques for remediation.

This special issue initially contains five articles, withplans for future publication of additional papers. The papersdeal with heavy metals and toxic organics, as well as soilacidification and the effects of military explosives on soil. Inthe paper “Occurrence of vanadium in Belgian and Europeanalluvial soils,” V. Cappuyns and E. Slabbinck bring attentionto the possible effects of V as a soil contaminant. They

document the nature and extent of V contamination inalluvial soils developed in three industrialized drainagebasins in Belgium, and from other areas in Europe. Theirresults suggest that the mobility of V is low, but neverthelessworthy of further investigation. B. V. Kjellerup et al. examinethe effects of biodecomposition by bacteria as a remedialtool for PCB contamination in the paper entitled “Spatialdistribution of PCB dechlorinating bacteria and activitiesin contaminated soil.” Their results support the use ofthe method for remediating sediments, whereas use incontaminated soils faces further challenges. In the papercalled “Acidification and nitrogen eutrophication of Austrianforest soils” R. Jandl et al. reevaluate the effects of acidicdeposition and nitrogen on forests soils. Interestingly, pHhas risen in the soils studied as a result of air pollutionmitigation and nitrate leaching into the groundwater is notfound to be a large-scale problem. The high levels of nitrogendeposition have actually led to an unexpected increase in theforest productivity. J. Pichtel reviews the effects of militaryexplosive wastes on soils in the article entitled “Distributionand fate of military explosives and propellants in soil: a review.”He shows that soils worldwide are contaminated by thechemically active components of explosives and propellants.These compounds undergo varying degrees of chemicaland biochemical transformation and appear to be commongroundwater contaminants. Thus, there appears to be anurgent need to identify and remove such hazards from con-taminated soils. In the paper “A critical evaluation of singleextractions from the SMT program to determine trace elementmobility in sediments,” V. Cappuyns compares two commonlyapplied single extraction procedures, ammonium-EDTA andacetic acid, for evaluating heavy metal contamination in

mailto:[email protected]

2 Applied and Environmental Soil Science

soils. The results underscore the difficulties of relating singleextractions to phytoavailability, and thus the need for furtherwork.

Fernando Jose GarbuioJeffrey L. Howard

Larissa Macedo dos Santos


Research Article

Acidification and Nitrogen Eutrophication ofAustrian Forest Soils

Robert Jandl,1 Stefan Smidt,1 Franz Mutsch,1 Alfred Furst,1

Harald Zechmeister,2 Heidi Bauer,3 and Thomas Dirnbock4

1 Federal Research and Training Centre for Forests, Natural Hazards and Landscape, Seckendorff-Gudent-Weg 8,1131 Vienna, Austria

2 University of Vienna, Rennweg 14, 1030 Vienna, Austria3 Division of Environmental and Process Analytics, Vienna University of Technology, Getreidemarkt 9/164,1061 Vienna, Austria

4 Environment Agency Austria, Spittelauer Lande 5, 1090 Vienna, Austria

Correspondence should be addressed to Robert Jandl, [email protected]

Received 2 December 2011; Revised 8 March 2012; Accepted 19 March 2012

Academic Editor: Larissa Macedo dos Santos

Copyright © 2012 Robert Jandl et al. This is an open access article distributed under the Creative Commons Attribution License,which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

We evaluated the effect of acidic deposition and nitrogen on Austrian forests soils. Until thirty years ago air pollution had led to soilacidification, and concerns on the future productivity of forests were raised. Elevated rates of nitrogen deposition were believed tocause nitrate leaching and imbalanced forest nutrition. We used data from a soil monitoring network to evaluate the trends andcurrent status of the pH and the C : N ratio of Austrian forest soils. Deposition measurements and nitrogen contents of Norwayspruce needles and mosses were used to assess the nitrogen supply. The pH values of soils have increased because of decreasingproton depositions caused by reduction of emissions. The C : N ratio of Austrian forest soils is widening. Despite high nitrogendeposition rates the increase in forest stand density and productivity has increased the nitrogen demand. The Austrian BioindicatorGrid shows that forest ecosystems are still deficient in nitrogen. Soils retain nitrogen efficiently, and nitrate leaching into thegroundwater is presently not a large-scale problem. The decline of soil acidity and the deposition of nitrogen together with climatechange effects will further increase the productivity of the forests until a limiting factor such as water scarcity becomes effective.

1. Introduction

Forests in central Europe provide manifold ecosystemservices. Besides their function for wood production andprotection, their soils represent an efficient filter and purifi-cation layer for water passing through, where they retaincarbon and nitrogen. The maintenance of carbon stocksin soils contributes to the mitigation of climate change.In addition, forest soils buffer acidic deposition. Over thelast three decades several factors have been identified whichcompromise the provision of these ecosystem services.

In the late 1970s, increasing evidence for forest declinearose [1]. The topic attracted a lot of public attention,followed by initiatives to reduce emissions and improveforest soils [2, 3]. In the mid 1980s, the nitrogen saturationhypothesis was brought up. Due to industrial processes

the nitrogen input to forests has reached unprecedentedlevels [4–6]. It was expected that the release of ammoniuminto soils is harmful due to the proton generation duringnitrification [7–9]. The formed nitrate is leached into thegroundwater, and as a consequence base cations are lost.The nitrogen saturation hypothesis was expanded to nutrientimbalances; these occur as a consequence of high growthrates, caused by excess nitrogen, when other nutrients arein short supply [10]. Following the general trend, Austrianforest soils are affected by acidic deposition. Many sites arecarrying a legacy of nutrient exploitation due to previousland-use practices. These sites were especially predisposedto acidification [11]. Soils formed on calcareous bedrockare less vulnerable. Historic forms of land use also haveaffected the nitrogen cycle. With respect to the nitrogensaturation hypothesis, it was expected that many forest soils


were critically depleted in nitrogen in the past. These forestsoils are, therefore, not easily oversaturated by incomingnitrogen, even when deposition loads exceed the actualnitrogen demand.

In Austria, the emission reduction between 1990 and2009 amounted to 72% for SO2, but only to 3% for NOx andto 4 % for NH3. Whereas mean annual NOx-concentrationsin Austrian nonurban areas have decreased over the last20 years, the Austrian limit value (30 μg NOx m−3 accord-ing to BGBl. 298/2001) as well as critical loads (10 to20 kg N ha−1 yr−1 depending on the respective ecosystem[12]) are often exceeded in the vicinity of roads [13, 14].Critical loads are a reference system defining upper limitsfor the deposition of elements that can safely be processedwithin the ecosystem. The deposition of elements/substancesbecomes problematic when the capacity is exceeded. Criticalloads are derived from the estimation of leaching rates fromsoils but also from the response of particularly vulnerableorganisms such as lichens, mosses, and vascular plants [15–18]. Air monitoring stations are irregularly installed andoften placed where high pollution is expected. Alternativemethods are necessary in order to provide a wider coverageof survey points. Bioindication with mosses is a useful tool toassess the nitrogen deposition [19–21].

(i) In this paper we show the long-term trend of soilpH in Austrian forest soils in order to evaluate theextent of the previous acidification and the recoveryof forest soils as a consequence of the decliningacidic deposition. We demonstrate the effect of theprolonged trend of nitrogen depositions on soilproperties.

(ii) We evaluate the effect on the basis of soil parametersand of receptor organisms, that is, the nutrition ofAustrian forest trees with nitrogen.

(iii) We compare the current deposition loads of protonsand inorganic nitrogen with previously establishedcritical loads. As a potential response to nitrogenenrichment, we present the temporal trend of thenitrogen nutrition of Austrian forests based on theAustrian Bioindicator Grid. As an indication for thespatial distribution of nitrogen supply, we presentnitrogen concentrations in mosses.

(iv) Pathways of nitrogen in forest ecosystems—for exam-ple, leaching of nitrate into the groundwater—will bedemonstrated by the example of two case studies.

2. Materials and Methods

2.1. Monitoring Networks. The evaluations are based onchemical data of soils and depositions and bioindication datafrom Austrian monitoring networks that are parts of severalindependently conducted transnational networks (Table 1).

2.2. Soils. The initial soil sampling was part of the ICP-Forests “Level I monitoring network” (http://www.icp-for-ests.net) and the second sampling was conducted within the

“BioSoil” project [22]. Data of pH values and C : N ratioswere selected from the comprehensive data pool and arediscussed here. The data set comprises 39 sites on calcareousbedrock and 100 sites on silicatic bedrock.

2.3. Depositions. The Austrian deposition monitoring gridfor forest ecosystem (ICP Forests programme) comprises20 plots representing the dominant forest types in Austria.The sites are located in clean-air regions and reflect thebackground level of atmospheric deposition. Furthermore,17 sample plots of the Federal deposition monitoringnetwork, located mainly in rural areas, were included.

Locations and analytical procedures of the Austriandeposition monitoring grid are described in [23], those ofthe Federal deposition monitoring network by [24]. Amonga series of ions the annual deposition rates of protons andtotal inorganic nitrogen (nitrate and ammonium) contentsfrom wet deposition were evaluated.

2.4. Bioindication with Spruce and Pine Needles. Within theAustrian Bioindicator Grid (http://bioindikatornetz.at; BIN)the pollutant input and the nutrition status of forests aremonitored. The BIN is organized on a regular grid andcomprises the content of main nutrients and micronutrientsin spruce and pine needles. Sampling was done annually inautumn. The nitrogen content of the foliage was used for thisevaluation. According to [25] concentrations <1.3% indicatenitrogen deficiency.

2.5. Bioindication with Mosses. The moss species Hylo-comium splendens, Pleurozium schreberi, Hypnum cupressi-forme, Abietinella abientina, and Scleropodium purum weresampled in 5-year intervals and analysed for nitrogen. Thesespecies have a similar morphological structure and a similarmechanism of nitrogen uptake. The objective of the mossanalysis was to obtain a proxy for the long-term inputof nitrogen when a comprehensive deposition monitoringsystem is not available. Here we use the content of nitrogen.

2.6. Case Studies at Intensive Investigating Plots. Input/out-put budgets of nitrogen were estimated for two long-termmonitoring plots in Austria.

The site Zobelboden belongs to the “Convention onLong-Range Transboundary Air Pollution” (CLRTAP)-Prog-ram and is located in the Northern limestone Alps (latitude:47◦50′ 30” N, longitude 14◦26′30′′; http://www.umwelt-bundesamt.at/im). The bedrock is formed by dolomite, thealtitude is 850–956 m, the average annual temperature is7.2◦C, and the annual rainfall ranges from 1500 to 1800 mm.The mixed forests are dominated by Norway spruce (Piceaabies) and beech (Fagus sylvatica). The site receives approx.20 kg N ha−1 yr−1 and, depending on the forest type, between21 and 27 kg N ha−1 yr−1 in the open field [26, 27].

The site Muhleggerkopfl is also located in the NorthernLimestone Alps (latitude 47◦34′ 50′′ N; longitude 11◦38′ 21′′

E) at 910 m a.s.l. on a north-northeast facing slope. Themean annual air temperature and precipitation are 6.8◦C

Applied and Environmental Soil Science 3

Table 1: Austrian monitoring networks regarding the evaluation of soil acidification and nitrogen saturation, nitrogen deposition andbioindication of nitrogen deposition, and nutrition status. The location of the sample points is shown in Figures 5 and 6.

Forest soils Deposition Spruce and pine needles Mosses

Project titleAustrian forest soilmonitoringsystem/BioSoil (EU)

Austrian deposition monitoringgrid for forest ecosystem (ICPForests); Federal depositionmonitoring network

Austrian Bioindicator Grid

Monitoring of atmosphericpollution by mosses in Austria(UNECE ICP monitoringprogramme)

Samplings 1987 and 2007 1996–2009; 1984–2009 1983–2010 1991, 1995, 2000, 2005, 2010

Number of sites 511 (1987); 139 (2007) 20/14 763 220

and 1563 mm, respectively. The experimental forest is 130-years old and dominated by Norway spruce. The site receivesapprox 20 kg N ha−1 yr−1 [28, 29].

2.7. Calculation of Critical Loads. The acidification ofsoils and the nitrogen saturation of forest ecosystems werecalculated based on the Austrian forest inventory [30], theAustrian inventory of forest soils [31], and Corine 2006 landcover data [32]. The sulphur, nitrogen, and cation depositionrates were estimated from time series modelled by the sources[33, 34].

The soil data were evaluated with the SAS statisticalpackage. For the change of soil pH and the C : N ratiobetween the first and the second soil inventory, a pairwisecomparison of means was done by a t-test. The data werestratified according to the geological substrate. The temporaltrend of the proton and nitrogen input between 1983 and2008 and the trend of the proportion of nitrogen deficientspruce and pine forests were evaluated by a linear regression.

3. Results

3.1. Soils. Austrian forest soils have consistently shown anincrease of the pH during two decades (Figure 1). Thestrongest increase occurred in the organic surface layer (FHhorizon) and the effect diminished with depth in the mineralsoil. The variability is quite large. The deacidification of soilswas stronger in soils on calcareous bedrock than on silicaticsoils. The large error bars are explained by the small-scalespatial variability of soil properties. The temporal changeof the pH is statistically significant in all depth layers ofthe calcareous soils and in the organic surface layer andthe upper mineral soil (0–10 cm) of the silicatic soils. As aresult of decreasing deposition of acidifying substances thecritical loads for acidification are exceeded at only 0.6% ofthe Austrian forest area.

The C : N ratio is an integrating indicator for theenrichment and the availability of nitrogen. A narrow C : Nratio indicates a high availability of nitrogen with severalconsequences such as an improved nitrogen supply for plantsand a higher soil microbial activity. According to Figure 2,the C : N ratio was slightly widening during the last 20 years.The effect is strongest in the forest floor material. Soils didnot respond differently among the 2 geological substrates.Contrary to the soil pH, the change in the C : N ratio is notstatistically significant, although the ratio has become wider.

1

0.9

0.8

0.7

0.6

0.5

0.4

0.3

0.2

0.1

0

−0.1

−0.2

Depth layer

pH Increase

Decrease

Carbonate soils

Silicate soils

1

0.9

0.8

0.7

0.6

0.5

0.4

0.3

0.2

0.1

0

−0.1

−0.2Litter layer 0–10 cm 10–20 cm 20–40 cm

Figure 1: pH change of Austrian forest soils between 1987 (old) and2007 (new) for different soil depths, differentiated for carbonate-free (silicatic, left bars) and carbonate-influenced (right bars) soils.

C/N Increase

Decrease

Depth layer

Carbonate soils

Silicate soils

Litter layer 0–10 cm 10–20 cm 20–40 cm

54.5

43.5

32.5

21.5

10.5

0−0.5−1

−1.5−2

−2.5

−3.5−3

54.543.532.521.510.50−0.5−1−1.5−2−2.5

−3.5−3

Figure 2: Change of the C/N ratio of Austrian forest soils between1987 (old) and 2007 (new) for different soil depths, differentiatedfor carbonate-free (silicatic, left bars) and carbonate-influenced(right bars) soils.


0.8

0.6

0.4

0.2

0

25

20

15

10

5

0

1985 1990 1995 2000 2005 2010

yr

N(k

g/h

a/yr

)H

(kg/

ha/

yr)

Figure 3: Annual input of protons (above) and inorganic nitrogen(below) in Austria.

3.2. Depositions. Long-term measurements of the wet protondeposition (free acidity) in Austria show a general decreaseover time. Between 1984 and 1993s the input of protonsdecreased substantially (between 39 and 99%) at five sitessituated at the northern Alpine rim and in the inner Alpinearea [24]. The highest inputs—up to 0.7 kg H+ ha−1 yr−1—were measured in the eastern part of Austria. The chemicalquality of the rain water varies widely in Austria dependingon the distance to emitters, and their exposure to importedair pollutants. In general, the pollutant loads correspondto the precipitation pattern so that higher deposition isparticularly found at the northern slopes of the Alps and tosome extent in southern Carinthia. The mean deposition ofprotons is 134 g H+ ha−1 yr−1 and decreases significantly by7 g H+ ha−1 yr−1 from 1984 to 2009 (P < 0.001; Figure 3,Table 2).

Nitrogen is entering the ecosystem as deposition ofammonium, nitrate, and organic nitrogen. The average totalinput of inorganic nitrogen is approximately 8 kg ha−1 yr−1.Figure 3 illustrates the considerable variation of nitrogendeposition, caused by the heterogeneity of sites. The variabil-ity is caused by the proximity of sites with high precipitationrates and high local emission and sites in the rain shadeof mountain ranges that receive very little deposition.During the observation period between 1996 and 2009the nitrogen input slightly, but significantly decreased. Thehighest inorganic nitrogen inputs (up to 22 kg ha−1 yr−1)

Table 2: Mean annual deposition rates of protons and nitrogen,mean annual changes, and significance of the changes at Austriameasurement sites. The sites are irregularly distributed.

ElementMean annual

deposition[kg ha−1 yr−1]

Mean annualchange [kg

ha−1]

Significanceof change

Numberof

samples

H 0.134 −0.007 ∗∗∗ 602

N (inorganic) 8.056 −0.146 ∗∗∗ 603

Table 3: critical loads (CL) for acidification (kg ha−1 yr−1) andeutrophication (kg ha−1 yr−1) of the Austrian forested area. CL-acidmax (S): critical load of S assuming an exclusion of N; CL-acidmax

(N): critical load of N assuming an exclusion of S; CL-eutnut (N): CLfor eutrophication according to the mass balance; CL-eutemp (N):empirical CL for eutrophication for Austrian forest types accordingto [26] ECE (2010).

25th percentile median 75th percentile

CL-acidmax (S) 2.384 2.776 6.242

CL-acidmax (N) 4.125 4.982 9.891

CL-eutnut (N) 10.1 11.0 11.9

CL-eutemp (N) 10–20

Percent area with an exceedance of the CL

CL-acid 0.6 %

CL-eut 94 %

were observed in the northern Alpine rim and in the easternpart of Austria, the lowest in the Inner Alps.

The critical loads for nitrogen eutrophication areexceeded at some sites. Average critical loads for acidificationare relatively high due to carbonate bedrock in many Austriaareas. Exceedance only occurred at 0.6% of the Austrianforest area. Critical loads for eutrophication based on massbalance are around 11 kg ha−1 yr−1. Nitrogen immobilisationincreases with altitude, whereas losses by nitrogen leachingdecrease with altitude. The critical loads for nitrogen varybetween 10 and 20 kg N ha−1 yr−1 over the Austrian terri-tory. In total, 94% of the Austrian forest area features adeposition exceeding the critical load for eutrophication. Theexceedance is 4.1 kg N ha−1 yr−1 on average and highest in thenorthern part of the Alps and in eastern Austria (Table 3).

3.3. Bioindication with Spruce and Pine Needles. The trend ofdecreasing nitrogen availability in soils, indicated by the C : Nratios, corresponds to the nitrogen deficiency in needles thatwas revealed by the Austrian Bioindicator Grid (Figure 4).According to the threshold values of 1.3% N, the area ofnitrogen deficient sites in Austrian forests has increased sincethe early 1980s. By contrast, the nitrogen supply has beendiminished and no evidence for nitrogen enrichment is givenbased on nitrogen contents of spruce needles.

The geographic distribution of the nutrition status hastotally changed during the last 30 years (Figure 5). Areasin the north and southeast of the country have been wellsupplied with nitrogen in the past and suffer currently from


0

10

20

30

40

50

60

7019

8319

8419

8519

8619

8719

8819

8919

9019

9119

9219

9319

9419

9519

9619

9719

9819

9920

0020

0120

0220

0320

0420

0520

0620

0720

0820

0920

10

(%)

Figure 4: Trend of the proportion of nitrogen deficient sites ofNorway spruce and Scots pine in Austrian forests from 1983 to 2010(Austrian Bioindicator Grid).

nitrogen deficiency. In the Central Alps the nitrogen supplyhas declined in the last 25 years.

3.4. Mosses. The distribution of the nitrogen content ofmosses can be taken as a surrogate for nitrogen depositionin areas not covered directly by trees or shrubs. As shownin Figure 6, nitrogen deposition has been elevated in denselypopulated areas, especially where the density of car trafficis high, where nitrogen was mainly deposited as nitrate.Furthermore, higher concentrations in the northern partof Austria are caused by intensive agriculture (e.g., cattleraising). In these areas ammonium was the main emissionsource [21]. The nitrogen content of mosses in the mountainregions is substantially lower.

3.5. Case Studies at Intensive Investigating Plots. Concerns onthe nitrogen saturation of forests invoked the implementa-tion of several case studies where the nitrate leaching fromforest ecosystems was quantified. At the two long-term mon-itoring sites Zobelboden and Muhleggerkopfl precipitation ishigh, so that even at low nitrogen concentrations in the raindeposition load is considerable. These sites are characterizedeither by mature spruce-dominated forests or mixed conifer-deciduous forests and a mosaic of deep chromic Cambisolsand shallow rendzic Leptosols, developed from the under-lying dolomite gravel. We assessed whether the soils werecurrently capable to retain the incoming nitrogen or whetherthey released nitrate into the karst aquifer. A repeated soilinventory at Zobelboden had indicated a slight increase inthe nitrogen pool of the organic surface layer and botanistsobserved an increase in the coverage of nitrogen demandingherbaceous plants and bryophytes [35, 36]. An estimationof the nitrogen losses by leaching showed no indication fornitrogen saturation. The nitrogen release in the aqueousphase, though quite high, was still dominated by specificevents such as strong snow melt, high amounts of litterfalldue to drought in the year 2003, or an increase of nitrogenavailability after a bark beetle attack that had taken outseveral trees and left some of the available nitrogen unutilized[36]. As a result, leaching of inorganic nitrogen can evenrise to 20 kg N ha−1 yr−1. The overall magnitude of nitrogen

leaching differs substantially between forests and soil types.The site in Tyrol also shows a remarkable nitrogen retentioncapacity. In all the years the nitrogen release in the aqueousphase was lower than the nitrogen input from deposition[37].

4. Discussion

The consistent increase of the soil pH within the last 20years gives evidence of strong changes in the Austrianforest ecosystems (Figure 1). The alkalinisation of forestsoils has also been shown consistently in Europe [38, 39].The decrease of soil acidity can partially be attributed tothe decrease of SO2 and NOx emissions which diminishedstress caused by acid deposition. However, effects of reducedemissions from industry are superimposed by the long-termrecovery of soils from negative impacts of former land usepractices [40]. Before mineral fertilisers became available forlow prices, shortly after World War II, forests commonlyserved as a source for organic material and nutrients foragriculture and have, therefore, been degraded [41]. Duringthe last 60 years, a recovery of forest soils has been observedin central Europe [42]. Therefore, the increase in soil pH isa combined response to both reduction of emissions fromindustrial sources and change in land use intensity.

The widening of the soil C : N ratio is an indicationfor a decreasing nitrogen availability in Austrian forest soils(Figure 2). Air pollution due to both imported long-rangetransport and local nitrogen emissions, caused considerablyhigher nitrogen deposition compared to the preindustrialamounts. Currently, the emissions are slightly declining. Inaddition the discontinuation of litter raking ensures that nolonger large quantities of nitrogen are removed from forestsoils. It was hypothesized that forest soils were approachingnitrogen saturation. However, we did not find evidence thatforest soils were accumulating nitrogen. A major reasonis the accentuated increase of the productivity of centralEuropean forests. The differentiation of simultaneous effectssuch as nitrogen enrichment, CO2 fertilisation, changedforest management and elongation of the growing seasonrevealed that nitrogen was the most important factor [43–45]. Moreover, the adverse effects of acidic deposition (SO2,NOx) are no longer effective on a large scale. Increasedgrowth rates of forests demand higher nitrogen supply fromsoils. The high productivity of forests is paralleled by lowernitrogen contents of the leaves and can only be supported bythe effective utilization of the soil nitrogen pool, as indicatedby the increase in nitrogen deficiency and the wider C : Nratio (Figures 2 and 4). Other factors such as increased soilorganic matter decomposition due to higher temperaturesand higher soil pH values [46] and improved types of forestmanagement are probably less important.

The change of the proton deposition is remarkable(Figure 3). The concern for the quality of the environmentled to legal measures obligating industry to efficiently reduceSO2 emissions. The implementation was successful due tothe well-defined industrial sources of SO2. The reduction ofnitrogen oxide emissions in order to counteract the effect


0 25 50 10 (km)

<1 µg/g <1 µg/g

1-1.1 µg/g

1.1-1.2 µg/g

1.1-1.2 µg/g

1.2-1.3 µg/g

1.3-1.4 µg/g

1.3-1.4 µg/g

1.5-1.6 µg/g

>1.7 µg/g

Federal province border>1.4 µg/g

Figure 5: Nitrogen content of moss tissue at 220 sampling sites in Austria following a method established within the framework ofUNECE/ICP vegetation. Dots represent measured concentrations in mosses; grey colours were derived from interpolation by means ofordinary kriging.

Bioindicator Grid 1983–1987Nitrogen deficiency

In 0 till 1 year

In 2 till 3 years

In 4 till 5 years

0 20 40 100 (km)60 800 20 40 100 (km)60 80

(a)

Bioindicator Grid 2006–2010Nitrogen deficiency

In 0 till 1 year

In 2 till 3 years

In 4 till 5 years

0 20 40 100 (km)60 80

(b)

Figure 6: Nitrogen nutrition of Austrian forest, based on the amounts of nitrogen deficiency in the period 1983–1987 (a) and 2006–2010 (b).

of nitrogen saturation was more difficult as the sources ofthese nitrogen compounds are more diffuse. The increaseof road traffic density led to increased nitrogen emissionrates that almost compensated the emission reductions inother sectors. Taking into account that SO2 emissions havedeclined and that NOx concentrations in the atmosphere didnot markedly decrease we conclude that sulphuric acid isnow of minor relevance for soil acidification, whereas nitricacid still supplies protons. Further sources of acidification arenitrification caused by high rates of NH4 deposition.

The decline of the nitrogen nutrition of forests (Figures 4and 5) is explained by observed forest management dynamics

during the last decades. Forest ecosystems in central Europehave been used less exploitative for several decades. Beingaware that closed nutrient cycles are relevant for ecosystems,their maintenance has been fully integrated into the conceptsof sustainable forest management. The nitrogen demandhas more strongly increased than the nitrogen supply inAustria due to afforestation of marginal agricultural landand the extended area of young forests with a high nitrogendemand [47]. Moreover, stand densities are increasing dueto economic constraints in terms of thinning operations.Political incentives towards the expansion of the productionof renewable energy are expected to increase the demand


for wood biomass in the future. A consequence would bethe increase of the management intensity and the intensifiedmobilisation of this resource [17, 48].

Trees are investing additional nitrogen in an increase ofthe needle biomass in order to achieve high productivityrates. The supply and the availability of nitrogen overan prolonged growing season as a consequence of risingtemperatures reduces the nitrogen content in photosynthet-ically active tissues. In conclusion, the increase of nitrogenavailability is not invested in nitrogen-enriched needles andleaves, but in an expansion of the tree. Fertilisation exper-iments demonstrated that nitrogen amendments led onlytransiently to higher concentrations in leaves, but primarilyreplenished the previously exhausted soil nitrogen pool[49].

The findings from needle analyses were confirmed bymoss analyses. In contrast to needles, for mosses theatmosphere is the main nitrogen source. The analysesindicate that the nitrogen content of mosses is a usefulintegrating parameter for the long-term supply of nitrogenfrom atmospheric sources [21, 35].

The case studies of nitrogen leaching did not give anyindication that the forest soils would lose their abilityto retain nitrogen. The nitrogen saturation hypothesisaccording to [9] is still considered to be valid. However,under natural conditions it could not be demonstrated thatmoderately elevated nitrogen depositions led to continu-ous and prolonged nitrogen leaching. Furthermore, meta-analyses of long-term studies revealed a number of factorsas prerequisites for losses due to nitrate leaching in additionto deposition [50]. In a comprehensive review [51], itwas demonstrated that nitrogen leaching is connected withecosystem disturbances. However, it is difficult to identifythe nitrogen enrichment of undisturbed forests, because theexcess nitrogen of a quantity of several kilograms per ha andyear dissipates in the large pool of several tons per ha.

Although model based mapping of critical loads has itsweaknesses [52], the widespread exceedance in Austria andelsewhere should attract more interest [53]. Global warmingprolongs the growing season and increases the productivityof forests. Consequently, the demand for nitrogen is expectedto increase. For regions with limited nitrogen supply, a“progressive nitrogen limitation” is expected [54, 55]. Globalwarming might, therefore, mitigate the problem of nitrogensaturation of forest ecosystems.

Forest disturbances are predicted to increase and havethe potential to mobilize reasonable amounts of N as shownby our case studies and elsewhere [27]. The uncertaintyregarding the identification of critical input rates cannot bea justification to ignore the challenge. From the viewpointof the chemical quality of water from forest ecosystems, it isimportant to know the limit of nitrogen retention capacity.

5. Conclusions

(i) In Austria, soils formed on noncalcareous bedrockare recovering from previous acidification due to therelease from exploitative forms of forest management

during centuries and acidic deposition in recenthistory.

(ii) The C : N ratio of Austrian forest soils is widening,despite a higher availability and quicker turnoverof nitrogen. The higher productivity of forests andchanges of the management intensity increase thenitrogen demand.

(iii) Nitrogen saturation is conceptually well understoodbut difficult to identify on the plot level when nitro-gen deposition is not massively exceeding criticalloads.

(iv) The nitrogen budgets of two intensive monitoringplots have shown that soils can retain high amountsof nitrogen, even when the input rate is exceeding thedemand by plants. Nitrate leaching is presently nota large-scale problem but could become one in thefuture when nitrogen emission is not reduced.

(v) The supply with nitrogen is worst among the mainnutrients nitrogen in Austrian forests.

(vi) Mosses are useful bioindicators for atmosphericnitrogen deposition as they are integrating the nitro-gen depositions of several years. Nitrogen concen-tration in well-defined moss species can be takenas proxies for nitrogen deposition in the respectivegrowth area.

References

[1] J. O. Reuss and D. W. Johnson, “Acid deposition and theacidification of soils and waters,” in Ecological Studies, vol. 59,Springer, 1986.

[2] R. Huttl, “Forest fertilisation: results from Germany, Franceand the nordic countries,” in Proceedings of the InternationalFertiliser Society, vol. 250, p. 40, 1986.

[3] R. Jandl, G. Glatzel, K. Katzensteiner, and O. Eckmullner,“Amelioration of magnesium deficiency in a Norway sprucestand (Picea abies) with calcined magnesite,” Water, Air, andSoil Pollution, vol. 125, no. 1–4, pp. 1–17, 2001.

[4] P. M. Vitousek, J. D. Aber, R. W. Howarth et al., “Human alter-ation of the global nitrogen cycle: sources and consequences,”Ecological Applications, vol. 7, no. 3, pp. 737–750, 1997.

[5] J. W. Erisman, “The European nitrogen problem in a globalperspective,” in The European Nitrogen Assessment, chapter 2,pp. 9–31, Cambridge University Press, 2011.

[6] M. A. Sutton, O. Oenema, J. W. Erisman, A. Leip, H. VanGrinsven, and W. Winiwarter, “Too much of a good thing,”Nature, vol. 472, no. 7342, pp. 159–161, 2011.

[7] B. Nihlgard, “The ammonium hypothesis. An additionalexplanation to the forest dieback in Europe,” Ambio, vol. 14,no. 1, pp. 2–8, 1985.

[8] J. D. Aber, “Nitrogen cycling and nitrogen saturation intemperate forest ecosystems,” Trends in Ecology and Evolution,vol. 7, no. 7, pp. 220–224, 1992.

[9] J. Aber, W. McDowell, K. Nadelhoffer et al., “Nitrogen satu-ration in temperate forest ecosystems: hypotheses revisited,”BioScience, vol. 48, no. 11, pp. 921–934, 1998.

[10] H. Van Oene, “Acid deposition and forest nutrient imbalances:a modeling approach,” Water, Air, and Soil Pollution, vol. 63,no. 1-2, pp. 33–50, 1992.


[11] R. Jandl, C. Alewell, and J. Prietzel, “Calcium loss in CentralEuropean forest soils,” Soil Science Society of America Journal,vol. 68, no. 2, pp. 588–595, 2004.

[12] World Health Organization, Air Quality Guidelines for Europe,European Series No. 91, WHO Regional Publications, 2ndedition, 2000.

[13] S. Smidt and E. Obersteiner, “10 Jahre Depositionsmessungim Rahmen des europaischen Waldschadensmonitorings,”Centralblatt fur das Gesamte Forstwesen, vol. 124, no. 2, pp. 83–104, 2007.

[14] S. Smidt, W. Spangl, and C. Nagl, “Trends von Schadstof-feintragen in osterreichischen Waldgebieten,” Centralblatt furdas Gesamte Forstwesen, vol. 127, no. 1, pp. 1–24, 2010.

[15] D. A. Burns, T. Blett, R. Haeuber, and L. H. Pardo, “Criticalloads as a policy tool for protecting ecosystems from the effectsof air pollutants,” Frontiers in Ecology and the Environment,vol. 6, no. 3, pp. 156–159, 2008.

[16] ECE, “Empirical critical loads and dose-response relation-ships,” 20th Session, Working Group on Effects ExecutiveBody for the Convention on Long-range TransboundaryAir Pollution, Economic Commission for Europe, Geneva,Switzerland, 2010.

[17] M. E. Fenn, S. Jovan, F. Yuan, L. Geiser, T. Meixner, and B. S.Gimeno, “Empirical and simulated critical loads for nitrogendeposition in California mixed conifer forests,” EnvironmentalPollution, vol. 155, no. 3, pp. 492–511, 2008.

[18] G. Velthof, “Nitrogen as a threat to European soil quality,”in The European Nitrogen Assessment, vol. 21, pp. 495–509,Cambridge University Press, 2011.

[19] M. E. Conti and G. Cecchetti, “Biological monitoring: lichensas bioindicators of air pollution assessment—a review,” Envi-ronmental Pollution, vol. 114, no. 3, pp. 471–492, 2001.

[20] B. A. Markert, A. M. Breure, and H. G. Zechmeister, Eds.,Bioindicators and Biomonitors. Principles, Concepts and Appli-cations, Elsevier, Amsterdam, The Netherlands, 2003.

[21] H. G. Zechmeister, A. Richter, S. Smidt et al., “Total nitrogencontent and δ15N signatures in moss tissue: indicative valuefor nitrogen deposition patterns and source allocation on anationwide scale,” Environmental Science and Technology, vol.42, no. 23, pp. 8661–8667, 2008.

[22] R. Hiederer and T. Durrant, “Evaluation of BioSoil demon-stration project—preliminary data analysis,” EUR 24258 EN,Office for Official Publications of the European Communities,Luxembourg, 2010.

[23] S. Smidt, “10 years of deposition monitoring within theEuropean forest monitoring programme,” BFW-Berichte, vol.138, 2007 (German).

[24] H. Puxbaum, V. Simeonov, and M. F. Kalina, “Ten yearstrends (1984–1993) in the precipitation chemistry in CentralAustria,” Atmospheric Environment, vol. 32, no. 2, pp. 193–202,1998.

[25] R. Huttl, “Forest fertilisation: results from Germany, Franceand the nordic countries,” in Proceedings of the InternationalFertiliser Society, vol. 250, p. 40, 1986.

[26] J. Kobler, W. J. Fitz, T. Dirnbock, and M. Mirtl, “Soil typeaffects migration pattern of airborne Pb and Cd under aspruce-beech forest of the UN-ECE integrated monitoring siteZobelboden, Austria,” Environmental Pollution, vol. 158, no. 3,pp. 849–854, 2010.

[27] G. Jost, T. Dirnbock, M. T. Grabner, and M. Mirtl, “Nitrogenleaching of two forest ecosystems in a Karst Watershed,” Water,Air and Soil Pollution, vol. 218, no. 1, pp. 633–649, 2011.

[28] S. Smidt, “Nitrogen fluxes in the Tyrolean Limestone Alps,”Environmental Science and Pollution Research International,no. 2, p. 52, 2002.

[29] A. Schindlbacher, S. Zechmeister-Boltenstern, and R. Jandl,“Carbon losses due to soil warming: do autotrophic and het-erotrophic soil respiration respond equally?” Global ChangeBiology, vol. 15, no. 4, pp. 901–913, 2009.

[30] K. Schieler and K. Schadauer, The Austrian Forest Inventory1992–96, Austrian Federal Office and Research Centre forForests, Vienna, Austria, 2011.

[31] Forstliche Bundesversuchsanstalt, “Austrian forest soil inven-tory,” in Mitteilungen der Forstlichen Bundesversuchsanstalt,vol. 168, Wien, 1992.

[32] CLC, “Corine Land Cover 2006,” EEA, Copenhagen, Den-mark, 2010, http://sia.eionet.europa.eu/CLC2006.

[33] EMEP, “EMEP/MSC-W modelled air concentrations anddepositions,” European Monitoring and Evaluation Pro-gramme, 2009, http://webdab.emep.int/Unified Model Re-sults/.

[34] M. van Loon, L. Tarrason, and M. Posch, “Modelling BaseCations in Europe,” EMEP/MSC-W & CCE Note x/2005-Draft1.2., 2005.

[35] H. G. Zechmeister, T. Dirnbock, K. Hulber, and M. Mirtl,“Assessing airborne pollution effects on bryophytes—lessonslearned through long-term integrated monitoring in Austria,”Environmental Pollution, vol. 147, no. 3, pp. 696–705, 2007.

[36] K. Hulber, T. Dirnbock, I. Kleinbauer et al., “Long-termimpacts of nitrogen and sulphur deposition on forest floorvegetation in the Northern limestone Alps, Austria,” AppliedVegetation Science, vol. 11, no. 3, pp. 395–404, 2008.

[37] R. Jandl, F. Herman, S. Smidt et al., “Modelling of nitrogendynamics in an Alpine forest ecosystem on calcareous soils: ascenario-based risk assessment under changing environmentalconditions,” Environmental Pollution, vol. 155, pp. 512–516,2008.

[38] D. Arrouays, P. H. Bellamy, and K. Paustian, “Soil inventoryand monitoring. Current issues and gaps,” European Journal ofSoil Science, vol. 60, no. 5, pp. 721–722, 2009.

[39] G. J. D. Kirk, P. H. Bellamy, and R. M. Lark, “Changes in soilpH across England and Wales in response to decreased aciddeposition,” Global Change Biology, vol. 16, no. 11, pp. 3111–3119, 2010.

[40] R. Jandl, F. Starlinger, M. Englisch, E. Herzberger, and E.Johann, “Long-term effects of a forest amelioration experi-ment,” Canadian Journal of Forest Research, vol. 32, no. 1, pp.120–128, 2002.

[41] J. Hofmeister, F. Oulehle, P. Kram, and J. Hruska, “Lossof nutrients due to litter raking compared to the effect ofacidic deposition in two spruce stands, Czech Republic,”Biogeochemistry, vol. 88, no. 2, pp. 139–151, 2008.

[42] J. Prietzel, E. Kolb, and K. E. Rehfuess, “Long-term monitoringof previously litter-raked pine forest sites in the provinceOberpfalz: changes in soil chemical properties and of forestnutrition,” Forstwissenschaftliches Centralblatt, vol. 116, pp.269–290, 1997 (German).

[43] K. Schadauer, “Growth trends in Austria,” in Growth Trends inEuropean Forests, H. Spiecker, K. Mielikainen, M. Kohl, and J.Skovsgaard, Eds., pp. 275–289, Springer, 1996.

[44] H. Spiecker, K. Mielikainen, M. Kohl, and J. Skovsgaard,“Growth Trends in Europe—Studies from 12 countries,” inGrowth Trends in European Forests, vol. 5, pp. 275–289,Springer, 1996.


[45] M. van Oijen and R. Jandl, “Nitrogen fluxes in two Norwayspruce stands in Austria: an analysis by means of process-basedmodeling,” Austrian Journal of Forest Research, vol. 121, pp.167–182, 2004.

[46] N. P. Lamersdorf and W. Borken, “Clean rain promotes fineroot growth and soil respiration in a Norway spruce forest,”Global Change Biology, vol. 10, no. 8, pp. 1351–1362, 2004.

[47] W. Russ, “Mehr Wald in Osterreich,” BFW-Praxisinformation,vol. 24, pp. 3–5, 2011.

[48] P. Fairley, “Introduction: next generation biofuels,” Nature,vol. 474, no. 7352, pp. S2–S5, 2011.

[49] C. Huber, W. Weis, and A. Gottlein, “Tree nutrition of Norwayspruce as modified by liming and experimental acidificationat the Hoglwald site, Germany, from 1982 to 2004,” Annals ofForest Science, vol. 63, no. 8, pp. 861–869, 2006.

[50] J. J. Rothwell, M. N. Futter, and N. B. Dise, “A classificationand regression tree model of controls on dissolved inorganicnitrogen leaching from European forests,” EnvironmentalPollution, vol. 156, no. 2, pp. 544–552, 2008.

[51] H. Van Miegroet and D. W. Johnson, “Feedbacks and syner-gism among biogeochemistry, basic ecology, and forest soilscience,” Forest Ecology and Management, vol. 258, no. 10, pp.2214–2223, 2009.

[52] R. A. Skeffington, “Quantifying uncertainty in critical loads. Aliterature review,” Water, Air, and Soil Pollution, vol. 169, no.1–4, pp. 3–24, 2006.

[53] M. Lorenz, H. D. Nagel, O. Granke, and P. Kraft, “Critical loadsand their exceedances at intensive forest monitoring sites inEurope,” Environmental Pollution, vol. 155, no. 3, pp. 426–435,2008.

[54] Y. Luo, B. Su, W. S. Currie et al., “Progressive nitrogen lim-itation of ecosystem responses to rising atmospheric carbondioxide,” BioScience, vol. 54, no. 8, pp. 731–739, 2004.

[55] D. W. Johnson, “Progressive N limitation in forests: reviewand implications for long-term responses to elevated CO2,”Ecology, vol. 87, pp. 64–75, 2006.


Research Article

Spatial Distribution of PCB Dechlorinating Bacteria andActivities in Contaminated Soil

Birthe V. Kjellerup,1 Piuly Paul,2 Upal Ghosh,2 Harold D. May,3 and Kevin R. Sowers4, 5

1 Department of Biological Sciences, Goucher College, 1021 Dulaney Valley Road, Baltimore, MD 21204, USA2 Department of Civil and Environmental Engineering, University of Maryland Baltimore County, 1000 Hilltop Circle, Baltimore,MD 21250, USA

3 Department of Microbiology & Immunology, Medical University of South Carolina, 171 Ashley Avenue, Charleston, SC 29425, USA4 Department of Marine Biotechnology, University of Maryland Baltimore County, 701 E. Pratt Street, Baltimore, MD 21202, USA5 Institute of Marine and Environmental Technology, Columbus Center, University of Maryland, 701 E. Pratt Street, Baltimore,MD 21202, USA

Correspondence should be addressed to Kevin R. Sowers, [email protected]

Received 2 December 2011; Accepted 28 February 2012

Academic Editor: Jeffrey L. Howard

Copyright © 2012 Birthe V. Kjellerup et al. This is an open access article distributed under the Creative Commons AttributionLicense, which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properlycited.

Soil samples contaminated with Aroclor 1260 were analyzed for microbial PCB dechlorination potential, which is the rate-limitingstep for complete PCB degradation. The average chlorines per biphenyl varied throughout the site suggesting that different rates ofin situ dechlorination had occurred over time. Analysis of PCB transforming (aerobic and anaerobic) microbial communities anddechlorinating potential revealed spatial heterogeneity of both putative PCB transforming phylotypes and dechlorination activity.Some soil samples inhibited PCB dechlorination in active sediment from Baltimore Harbor indicating that metal or organiccocontaminants might cause the observed heterogeneity of in situ dechlorination. Bioaugmentation of soil samples contaminatedwith PCBs ranging from 4.6 to 265 ppm with a pure culture of the PCB dechlorinating bacterium Dehalobium chlorocoercia DF-1 also yielded heterologous results with significant dechlorination of weathered PCBs observed in one location. The detection ofindigenous PCB dehalorespiring activity combined with the detection of putative dechlorinating bacteria and biphenyl dioxygenasegenes in the soil aggregates suggests that the potential exists for complete mineralization of PCBs in soils. However, in contrast tosediments, the heterologous distribution of microorganisms, PCBs, and inhibitory cocontaminants is a significant challenge forthe development of in situ microbial treatment of PCB impacted soils.

1. Introduction

Polychlorinated biphenyls (PCBs) are persistent organicpollutants that are still present in the environment despitea U.S. production ban in 1976 [1]. Prior to this, commercialmixtures of PCBs (trade name Aroclor in the U.S.) were usedfor a range of industrial applications such as high-voltagetransformers, insulating materials, and hydraulic liquids [2,3]. PCBs are hydrophobic with a high affinity for adsorptionto soil particles and for bioaccumulation in lipids causinghepato- and immunotoxicity, carcinogenesis, and affectingendocrine organs and reproduction in humans [4, 5] andanimals [6]. Removal of PCBs from impacted sites has,therefore, been a regulatory priority for several decades [7].

Soils contaminated with PCBs can be found worldwideas a result of industrial activity [8]. However, large het-erogeneities in contaminant concentration and microbialpopulations were observed, when the total concentrationof PCBs and other contaminants were evaluated on bothmacro- and microscales [9, 10]. In some cases, the reason forthe heterogeneity was caused by the source of contaminationsuch as organic contaminants from an aluminum plant,where there was a decreasing contaminant gradient awayfrom the plant [11]. Even in cases without a specific source,the physical and environmental conditions in soil involved information of soil aggregates impacted the spatial distributionof PCBs leading to heterogeneity of microbial activity [12,13]. Despite the physical heterogeneities in soil, prior studies


have shown that this variability is not a crucial factor forthe composition of the microbial soil communities [14].Previous reports suggested that sites contaminated withweathered PCBs are recalcitrant to microbial dechlorination[15] due to sequestering of PCB molecules in the pores ofthe soil particles [16]. Desorption of PCBs depends on thepartitioning of PCB molecule between the pore water andthe associated organic matter as well as the diffusion of themolecules. It has been suggested that formation of strongbonds between the PCB molecule and the soil particles couldcause desorption resistance [16].

Complete microbial degradation of PCBs requires anaer-obic reductive dechlorination of extensively chlorinatedcongeners followed by subsequent aerobic cleavage of thebiphenyl ring and mineralization of the less extensivelychlorinated congeners [17, 18]. Several anaerobic bacteriawithin the Chloroflexi have been confirmed to have PCBdechlorinating activity including Dehalococcoides ethano-genes, [19], Dehalococcoides sp. CBDB1 [20], Dehalobiumchlorocoercia DF-1, bacterium o-17 [21], and phylotypes SF-1 and DH-10 [22]. In contrast, aerobic PCB degradation canbe performed by many bacterial species such as Burkholderiaxenovorans strain LB400 [23], Rhodococcus sp. strain RHA1[24], and bacteria utilizing biphenyls are ubiquitously dis-tributed in the environment [25]. Biphenyl 2,3-dioxygenasesare considered the key enzymes in the oxidative pathwayfor PCB degradation, where bphA1 can degrade specificPCB congeners [26] and bphC is involved in extradiol metacleavage [27]. In the soil environment, bioaugmentationwith aerobic PCB degrading bacteria has successfully beenapplied [28, 29], whereas anaerobic bioaugmentation haspreviously only shown success in sequential anaerobic-aerobic treatment in granular sludge and a mixture ofsediment and soil [30].

In this study, the spatial distribution of putative aerobicand anaerobic bacteria involved in PCB transformation wasinvestigated for the first time in PCB-contaminated soil. Theindigenous anaerobic dechlorination potential was mappedspatially throughout the site using a fixed grid established forprevious sampling events. In addition, the effect of anaerobicbioaugmentation with a PCB dechlorinating microorganismwas evaluated.

2. Materials and Methods

2.1. Bacterial Cultures for Bioaugmentation. Dehalobiumchlorocoercia DF-1 isolated from sediments in CharlestonHarbor was maintained in the laboratory as descried previ-ously [31]. D. chlorocoercia dechlorinates doubly flanked inthe para or meta positions.

2.2. Sample Collection. Samples were collected from astormwater drainage ditch in Mechanicsburg, PA ranging from itssource at 40◦13′46.16

′′N, 76◦59′38.51

′′W to approximately

730 meters downstream to 40◦14′06.60′′

N, 76◦59′32.50′′

W.The open drainage ditch, which collects storm water runofffrom the Naval Support Activities base and surrounding off-base properties, extends approximately 2.4 kilometers from

Meb19Meb20

Meb1Meb2

Meb10Meb11Meb12 Meb13

Meb14Meb15

Meb16Meb17Meb18 Meb3

Meb4Meb5 Meb6

Meb7 Meb8365 m

N

W E

S

365 m

Figure 1: Sampling locations in the drainage ditch.

its origin to its confluence with Trindle Spring Run. Samplescontaminated with different levels of PCBs were collectedwith a 60 cm by 5 cm (OD) core sampler from the top 30 cmof soil. In total, 19 soil samples were collected (Figure 1) andstored anaerobically in sealed glass jars at 4◦C in the dark andprocessed 1–3 weeks later.

2.3. Extraction and Analysis of PCBs from Soil Samples andMicrocosms. The total concentration of PCBs in the sampleswas measured in sacrificed samples as described previously[32]. Briefly, the samples were weighed prior to extraction,the overlying water was separated from the soil by decanting,and the remaining sample was mixed with anhydrous sodiumsulfate. Following the addition of the surrogates PCB-14,PCB-65, and PCB-166, the overlying water samples wereshaken in hexane, while the solid samples were extracted bysonication in 25 mL of 1 : 1 acetone : hexane for six minutesfollowing EPA method 3550B. The extracts were then pooledand solvent exchange was performed to replace the solventswith hexane. PCB cleanup was based on EPA SW846Methods 3660B (activated copper treatment), and 3630C(silica gel treatment). The volume obtained after cleanup wasdiluted according to QA protocols, and the PCB congenerswere analyzed using an Agilent 6890 Gas Chromatographequipped with microelectron capture detector.

2.4. Microbial Activity Assays. A microbial dechlorinationactivity assay was used to determine the potential dechlori-nation activity of indigenous microbial communities in thesoil samples as described previously [33]. Briefly, 10 mL oflow-sulfate mineral F-medium [34] was prepared anaero-bically and the congener 2,3,4,5,6-CB (AccuStandard, CT)was added to a final concentration of 50 ppm in acetone(10 µL added). Triplicate cultures were inoculated with 4.0 gof soil (wet weight). Negative controls were prepared byautoclaving twice for 20 min at 121◦C one and three daysafter inoculation prior to adding PCB. The cultures wereincubated at 30◦C in the dark and 1 mL subsamples werecollected for PCB analysis in an anaerobic glove box after 0,52, 75, 105, 130, 163, and 200 days.

Inhibition of the dechlorination activity was evaluatedwith a modification of the activity assay described above.Sediment from Baltimore Harbor (2 g) with confirmeddechlorination activity was mixed with the test soil sample


(2 g) in E-Cl medium and analyzed for PCB dechlorination.Sampling for PCB analysis was performed after 0, 50, 75, 100,150, and 214 days.

2.5. DNA Extraction. DNA was extracted (in triplicate)by transferring 0.75 g of soil to a sterile microcentrifugetube containing approximately 1 g of 0.1 mm Zirconia/SilicaBeads (BioSpec Products, Inc. OK) followed by additionof 200 µL of 1x TE buffer, 150 µL of phosphate buffer pH8.0 (0.12 M) and 150 µL 1x TS-SDS buffer. The sampleand buffers were mixed by hand shaking prior to 30 sof bead beating at speed “4.5” using a FastPrep120 (Q-Biogene, CA). Nucleic acid extraction and purificationwere performed using a phenol/chloroform-based protocoldescribed previously [35].

2.6. Detection and Enumeration of Putative PCB Dechlo-rinating Bacteria. Enumeration of putative dechlorinatingbacteria in the soil samples was performed in triplicateby a competitive PCR assay as described previously [33]using the 348F-884R 16S rRNA gene primer set. To confirmthat potential PCR inhibitors were not coextracted, allsamples were tested in parallel with the universal 16SrRNA gene primer set 341F/907R [36, 37]. All samplesfrom the site showed positive results with the universalprimer set indicating that the assay would detect putativedechlorinating phylotypes in abundances greater than thedetection limit of 2× 102 gene copies µL−1. The enumerated16S rRNA genes copies from the cPCR assay were normalizedto the dry weight content of the soil sample. One 16SrRNA gene copy per cell was assumed based on the genomesequences of Dehalococcoides ethenogenes [38] and CBDB1[39]. DNA extraction efficiency was tested by extractingdifferent amounts of homogenized soil and measuring theDNA concentration on a spectrophotometer at 280 nm. Theresults showed linear extraction efficiency in the range from0–4 g dry wt of soil (data not shown).

2.7. Detection of Potential Aerobic PCB Degrading Bacteria.Putative PCB degrading aerobic bacteria were detected byPCR amplification of the functional genes bphA and bphC,which encode the PCB transforming enzymes biphenyldioxygenase and 2,3-dihydroxybiphenyl 1,2-dioxygenase,respectively. Detection of bphA was performed with primerset bphA40-F/bphA50-R [26], whereas detection of bphCwas performed with primer set P42D-F/P43U-R [27].PCR was conducted in 50 µL reaction volumes using thefollowing GeneAmp reagents (Applied Biosystems, FosterCity, CA): 10 mM Tris-HCl, 75 mM KCl, 0.2 mM of eachdNTP in a mix, 1.5 mM MgCl2, 2.5 units of AmpliTaq DNAPolymerase, 50 pM of each primer, 1 µL of DNA template,and 34.5 µL of nuclease free water. Amplification of the bphAfragment was performed as follows (40 cycles): denaturationat 45◦C for 60 s, primer annealing at 31–49◦C for 3 min,elongation at 72◦C for 4 min, and a final holding step at4◦C. For amplification of the bphC fragment (35 cycles), thefollowing conditions were applied: denaturation at 95◦C for30 s, primer annealing at 35◦C for 60 s, elongation at 72◦C

for 3 min, a final extension step at 72◦C for 10 min, and afinal holding step at 4◦C. PCR products of the correct lengthwere confirmed by electrophoresis using a 1.5% agarose gel.Burkholderia xenovorans LB400 [23] was used for verificationof the PCR protocols for bphA and bphC and as the positivecontrol during PCR amplification.

2.8. Bioaugmentation in Microcosms. In total, 4 g of soil(wet weight) was inoculated into triplicate 10 mL volumesof anaerobically prepared mineral medium [41] in 25 mLanaerobe tubes sealed under N2-CO2 (80 : 20) with Teflonsepta. No electron donors were added except for thecomponents in the medium (0.0125% w/v cysteine) andresidual hydrogen (≤5% v/v) present in the atmosphere ofthe anaerobic glove box used for inoculating and samplingof the microcosms. D. chlorocoercia DF-1 was grown toapproximately 107 cells per mL with PCE, which was purgedwith N2/CO2 to remove residual PCE and chloroetheneproducts that could result in “priming” of dechlorinationactivity. 2 mL of culture were inoculated into the soilmicrocosms. Nonbioaugmented controls included mediumand soil containing indigenous microorganisms without DF-1. Controls for abiotic activity containing medium, soil con-taining indigenous microorganisms, and DF-1 were sterilizedby sequential autoclaving for 1 hour on days 0, 2, and 4.

2.9. Microbial Community Analysis. Community analyseswere performed by denaturing HPLC (DHPLC) using aWAVE 3500 HT system (Transgenomic, Omaha, NE) and the16S rRNA gene primers 348F/884R as described previously[33]. The 16S rRNA gene fragments were analyzed in20 µL injection volumes and the fractions were collectedin 96 well plates (Biorad, Hercules, CA). Fractions weredried using a Savant SpeedVac system (Thermo ElectronCorporation, Waltham, MA) followed by dissolution in15 µL nuclease-free water. Each DHPLC fraction wassequenced in both the 5′ and 3′ direction with 250 pMof primer 348F or 884R, respectively, in 5% DMSO toreduce effects from potential secondary structure using theBigDye Terminator v3.1 (Applied Biosystems, Foster City,CA) kit per the manufacturer’s instructions and sequencedon an ABI 3130 XL automated capillary DNA sequencer(Applied Biosystems, CA) as previously described [33]. The16S rRNA gene sequences and submitted gene sequencesobtained from NCBI (http://ncbi.nlm.nih.gov/BLAST)were compiled and aligned using the automatic nucleicacid aligner in the BioEdit sequence alignment editor.A total of 13 sequences containing approximately 530nucleotides were unambiguously aligned and used forcalculation of trees by the neighbor joining and FITCHapproaches using default settings in the PHYLIP software(http://evolution.genetics.washington.edu/phylip.html).Bootstrap analyses (1000 replicates) were performed usingthe PHYLIP package.

2.10. Nucleotide Accession Numbers. The Genbank accessionnumbers for the 16S rRNA sequences reported in this paperare JF412634-JF412646.


Table 1: Analyses of the dry matter content, the total PCB concentration, the average number and percentage of chlorines less than 4 and6, respectively, and in soil samples and commercial Aroclor mixtures [40]. a,b,crefer to the homolog distribution profiles, where ais the mosttypical distribution, bis the most dechlorinated, and cis the least dechlorinated among the soil analyzed soil samples. Sample Meb9 was notcollected.

Location Dry matter (mg/g) PCB conc. (ppm) Average no. of chlorines ≤6 chlorines (%) ≤4 chlorines (%)

Meb1 18.92 34.8 6.52 52.1 0.4

Meb2a 17.54 264.6 6.47 53.6 0.5

Meb3 13.71 12.5 6.51 52.5 0.5

Meb4 28.25 36.6 6.54 50.7 0.4

Meb5 23.17 15.1 6.54 50.5 0.4

Meb6 26.61 3.1 6.43 54.7 1.7

Meb7 16.71 5.6 6.50 52.6 0.2

Meb8 23.71 45.3 6.54 50.5 0.5

Meb10b 15.91 4.6 6.36 55.9 4.3

Meb11c 26.68 2.3 6.61 45.8 1.0

Meb12 33.61 8.2 6.48 52.3 2.4

Meb13 13.55 12.1 6.47 51.8 3.2

Meb14 27.34 12.3 6.47 52.3 2.6

Meb15 29.22 35.6 6.54 47.1 2.6

Meb16 15.20 9.1 6.48 50.9 2.7

Meb17 25.10 17.6 6.52 50.9 2.0

Meb18 36.72 11.3 6.50 50.9 2.7

Meb19 20.31 20.1 6.55 48.7 1.9

Meb20 18.60 72.7 6.54 51.7 0.3

Aroclor 1016 — — 3.04 100.0 100.0

Aroclor 1232 — — 2.41 100.0 94.1

Aroclor 1242 — — 3.31 100.0 92.3

Aroclor 1248 — — 3.97 100.0 82.2

Aroclor 1254 — — 5.15 96.0 18.9

Aroclor 1260 — — 6.39 56.7 0.6

3. Results

3.1. Physical and Chemical Characterization of Soil Samples.Soil samples from 19 locations were collected in the centerof the storm water drainage ditch bed and on each ofthe flanking banks at incremental distances from the inletto approximately 365 m downstream in addition to fivelocations between 365 and 730 m downstream from theinlet (Figure 1). Soil samples collected from the banks wereall moist, whereas both dry and moist soil samples werecollected from the center of the ditch. Soil was not collectedfrom the center of the ditch at the inlet because only 1-2 cm of soil was present above the bedrock due to erosionfrom heavy storm water events, but this layer increased indepth with distance from the inlet. Soil samples collectedmore than 365 m downstream were submerged in water atthe time of sampling. The total PCB concentration at the 19sampling sites ranged from 2.3–265 ppm with an average of32.8 ppm (±58.9 ppm), (Table 1). The highest concentrationwas detected at the east bank nearest to the inlet (Meb2),whereas the lowest concentration was detected at the westbank sampling location Meb11. The average number ofchlorines at each site ranged from 6.36–6.61, with a total

average of 6.50 (±0.053), which is similar to the average of 6.4observed for Aroclor 1260 (Table 1). The distribution of PCBhomologs was similar to A1260 and in prior studies A1260was reported to be the only Aroclor detected at this site[42]. In some of the samples, higher concentrations of tetrachlorinated congeners were observed (Table 1). The elevatedlevels of tetrachlorinated congeners detected did not coelutewith any known chlorinated non-PCB compounds includingpesticides and sulfur indicating that they were likely PCBs.The results suggest that in situ PCB dechlorination occurredto some extent in these samples.

3.2. Characterization of Indigenous Microbial Communities.Bacteria were detected in 18 of 19 samples using universal16S rRNA bacterial primers (Table 2). In contrast, putativeanaerobic dechlorinating and putative aerobic degradingbacteria were detected in only ten samples using primersspecific for dechlorinating Chloroflexi and bphA/bphC genes,respectively (Table 2). At seven locations, both putativeaerobic and anaerobic PCB transforming bacteria weredetected in the same samples. The numbers of putativeanaerobic dechlorinating bacteria ranged from 5·103 to


Table 2: Analyses of total bacteria, putative PCB dechlorinating and PCB degrading bacteria and the number of putative dechlorinatingbacteria in soil samples. Sample Meb9 was not collected.

LocationPresence/absence of

Bacteria Dechlorinating bacteria Aerobic PCB degradersNo. of dechlorinating bacteria

(16S copies·g soil−1)

Meb1 + − − —

Meb2 + − − —

Meb3 + + + <DLa

Meb4 + + + 3 · 104 ± 2.8 · 104

Meb5 + + + <DL

Meb6 + + + 3 · 106 ± 2.8 · 106

Meb7 + + + 4 · 105 ± 2.3 · 105

Meb8 + + − 3 · 105 ± 2.8 · 105

Meb10 + − + —

Meb11 + − − —

Meb12 + − − —

Meb13 + − + —

Meb14 + + + <DL

Meb15 + + + <DL

Meb16 + - + —

Meb17 + + − <DL

Meb18 + + − 3 · 103 ± 2.8 · 103

Meb19 + − − —

Meb20 + − − —

Mean values and standard deviations ± are given (n = 3).aDL: detection limit ≥102 16S copies/g wet soil.

5·106 dechlorinating bacteria g−1 soil, whereas the numbersat several locations were below the detection limit forenumeration, but presence was detected. The number ofputative aerobic degrading bacteria was not determined forthis phylogenetically diverse group of bacteria.

The population of putative dechlorinating bacteria wasexamined by DHPLC in three samples where 16S rRNAgene copies were most abundant (Meb6, Meb7, and Meb8)and compared to the sequences of known PCB dechlori-nating bacteria within the Chloroflexi group (Figure 2). Fiveof the 13 identified phylotypes grouped closely togetherwithin the clade of known Dehalococcoides sp. and theremaining eight phylotypes grouped within the broader o-17/DF-1 Chloroflexi group. No trend was observed betweenthe location of the samples and the identified phylotypes,since phylotypes from the three locations were locatedthroughout the phylogenetic tree. This observation indicatesthe population of indigenous phylotypes was heterogeneousthroughout the site.

3.3. Dechlorination Activities by Indigenous Communities.The indigenous dechlorination potential was examined insoil samples using 2,3,4,5,6-CB as a surrogate, which issaturated with chlorines on one biphenyl ring. Activity wasdetected in six of the 19 samples (Table 3). The highestdechlorination rate (6.93 ± 5.33 · 10−3 mol% 2,3,4,5,6-CB× day−1) was measured in Meb16 and the lowest (0.93

± 1.70 · 10−3 mol% 2,3,4,5,6-CB × day−1) in Meb17.The observed products of the dechlorination of 2,3,4,5,6-CB were predominantly 2,3,4,5-CB/2,3,5,6-CB, but 2,3,6-CB, and 2,6-CB were detected in sample Meb10. Theremaining 13 samples did not show any dechlorinationactivity after 200 days of incubation despite detection ofputative dechlorinating bacteria by cPCR in some of theinactive samples.

3.4. Effects of Indigenous Contaminants on Activity. Theabsence of detectable dechlorination activity in some ofthe samples despite the detection of putative dechlorinatingphylotypes was further examined to determine whether highconcentrations of PCBs or other contaminants inhibitedactivity. Soil from inactive samples was mixed with bio-catalytically active sediment from Baltimore Harbor, MD,which transforms 2,3,4,5,6-CB to 2,4,6-CB via reductivedechlorination at the two meta positions. For all testedsamples including sediment from Baltimore Harbor, thelag phase was less than 50 days and the terminal plateauranged between 6 and 46 mol% 2,3,4,5,6-CB (Table 4). Thedechlorination rates varied from 7–12 · 10−3 mol% 2,3,4,5,6-CB day−1, where the most active site had approximately25% greater activity than observed in Baltimore Harborsediment alone. This soil sample (Meb7) was collectedfurthest away from the inlet of the drainage ditch, had thehighest number of putative dechlorinating bacteria, and had


Deep subsurface paleosol clone 0.6. (AF005745)Deep subsurface paleosol clone 1.4. (AF005748)Deep subsurface paleosol clone 3.93 (AF005750)

Dehalococcoides ethenogenes 195 (AF004928)DCEH2 (AJ249262)Dehalococcoides sp. clone DHC-vic (AF3888550)Dehalococcoides sp. clone DHC-dll (AF3888536)

Dehalococcoides sp. FL2 (AF357918)DEH10 (DQ021869)Dehalococcoides sp. CBDB1 (AF230641)

Dehalococcoides sp. BAV1 (AY165308)Dehalococcoides sp. JN18 V108 B (EF059530)

Dehalococcoides sp. clone Fa3/FE2 (AY553936)

Dehalococcoides sp. clone PMVC2 (DQ833296)Putative dechlorinating Chloroflexi DHPLC fraction AN4 (EF682186)

Putative dechlorinating Chloroflexi DHPLC fraction GR1 (EF682195)Putative dechlorinating Chloroflexi DHPLC fraction BU3 (EF682191)Putative dechlorinating Chloroflexi DHPLC fraction BU2 (EF682190)

Putative dechlorinating Chloroflexi DHPLC fraction GR2 (EF682196)

SF1 (DQ021870)DF-1 (AF393781)

Chloroflexi clone pltb-vmat-61 (AB294962)Chloroflexi clone S2-69 (EF491385)

Chloroflexi clone 92 36 (AY534100)

−100

Dehalococcoides sp. JN18 A96 B∗ (DQ168647)

Putative dechlorinating Chloroflexi DHPLC fraction AN5 (EF682187)

Oscillochloris trichoides (AF093427)

t fH fH

MEB11 (JF412644)

MEB6 (JF412639)

MEB10 (JF412643)MEB5 (JF412638)MEB13 (JF412646)

MEB4 (JF412637)MEB3 (JF412636)

MEB1 (JF412634)MEB2 (JF412635)

MEB9 (JF412642)

MEB8 (JF412641)MEB12 (JF412645)

MEB7 (JF412640)

Figure 2: Phylogenetic tree showing the relationships between the dominant phylotypes identified in sediment samples Meb6, Meb7, andMeb8 (bold) and the closest dechlorinating species within the dechlorinating Chloroflexi group. Phylotypes from sample Meb6, Meb7 andMeb8 were located throughout the phylogenetic tree and did not show any relationship between the sample origins their phylogeneticrelationship. Accession numbers are indicated in parentheses. The tree was calculated by the neighbor joining method and supported byFITCH [43]. The scale bar indicates 10 substitutions per 100 nucleotide positions.

a high water content compared to the rest of the samples.The dechlorination rates for the mixed samples were allhigher or at levels similar to those observed in the activityassays (Table 3). In all samples tested, the end product ofdechlorination was 2,4,6-CB, which is characteristic of BHactivity rather than the tetrachlorinated congener productsdetected predominantly in indigenous ditch microcosms,indicating that the indigenous BH dechlorinating activitieswere dominant. A statistical evaluation (Student’s t-test,P < 0.05) showed that statistically significant inhibition wasfound for only samples Meb6 and Meb18 (dechlorinationplateau) and Meb17 (dechlorination rate and plateau).

3.5. Effect of Bioaugmentation. Bioaugmentation with apure culture of the PCB dechlorinating bacterium DF-1was tested in microcosms with soil from three locationscontaining approximately 5 ppm (Meb10), 73 ppm (Meb20),and 265 ppm (Meb2) total weathered PCBs. These locationswere selected based on the different concentrations ofPCBs. The specific dechlorination of only doubly flankedchlorines by this organism could be readily distinguishedfrom the dechlorination of single flanked and unflanked

chlorines by the indigenous populations. Controls withindigenous populations without DF-1 and autoclavedcontrols showed no significant dechlorination. In thebioaugmented microcosms, significant dechlorination wasobserved for Meb10 (5 ppm), (Table 5). In addition to DF-1,a highly enriched culture of indigenous PCB dechlorinatingmicroorganisms from MEB10 enriched with 2,3,4,5,6-PCBwas also used to bioaugment Meb2, but no significanteffect on dechlorination was observed. Evaluation of thechanges in homolog distributions indicated that DF-1 wasactive in Meb10 (Figure 3), with significant reductions ofhepta- and octachlorinated congeners and significantincreases of tetra- and pentachlorinated congeners.The congeners that were significantly reduced were22′33′45′6 (PCB-175), 22′33′55′66′/22′33′44′6/233′44′5(PCB-202/171/156), 22′33′455′ (PCB-172), 22′344′55′

(PCB-180), 22′33′44′5/233′44′56 (PCB-170/190), and22′344′55′6/22′33′44′56′ (PCB-203/196). All congenersthat were dechlorinated significantly more than the controlcontained both double-flanked meta and para positionedcongeners with the exception of PCB-175 and PCB-171 thatonly contained a double-flanked meta positioned chlorine


Table 3: The potential dechlorination activity of the soil samples in microcosms. Locations not listed here did not show any dechlorinationactivity.

LocationMax rate·10−3a

(mol% 2,3,4,5,6-CB × day−1)Lag phase (d)

PCB 116 remaining after 200 days(mol% 2,3,4,5,6-CB)

Dechlorination products

Meb10 4.2 (0.7) 0–50 46.0 (7.0)2,3,4,6-CB/2,3,5,6-CBb

2,3,6-CB, 2,6-CBc

Meb12 4.2 (4.1) 0–130 82.3 (5.8) 2,3,4,6-CB/2,3,5,6-CB

Meb13 3.4 (3.8) 0–150 78.0 (15.6) 2,3,4,6-CB/2,3,5,6-CB

Meb16 6.9 (5.3) 0–105 58.7 (27.6) 2,3,4,6-CB/2,3,5,6-CB

Meb17 0.9 (1.7) 0–200 89.7 (18.2) 2,3,4,6-CB/2,3,5,6-CB

Meb19 3.0 (5.0) 0–200 68.3 (49.7) 2,3,4,6-CB/2,3,5,6-CB

Negative control 0.1 — 98 —aMean values are given and figures in brackets are standard deviations (n = 3).

bThe products 2,3,4,6-CB/2,3,5,6-CB coeluted during GC analysis.cThis product was detected in one of three replicate cultures <5 mol%.

Table 4: Effect of inhibition on the dechlorination activity in microcosms using selected soil and sediment samples mixed 1 : 1 with activelydechlorinating sediment from Baltimore Harbor. Inhibition was determined (P < 0.05) for the dechlorination rate and terminal plateau. Lagphase is defined as not showing activity that is different from the autoclaved negative control.

Location Lag phase (d)PCB 116 remaining after 200 days(mol% 2,3,4,5,6-CB)

Dechlorination rate · 10−3

(% 2,3,4,5,6-CB day−1)End product

Inhibition (P < 0.05)

Rate Plateau

Baltimore Harbor 0–50 5.7± 4.5 9.1± 2.2b 2,4,6-CBa — —

Meb6 0–50 31.3± 15.9 7.3± 1.0 2,4,6-CB No Yes

Meb7 0–50 10.7± 3.5 12.4± 1.1 2,4,6-CB No No

Meb8 0–50 21.0± 19.2 8.4± 3.7 2,4,6-CB No No

Meb17 0–50 46.3± 6.5 7.5± 2.2 2,4,6-CB Yes Yes

Meb18 0–50 45.7± 10.1 10.5± 2.2 2,4,6-CB No YesaOne tube showed additional terminal products such as 2,4-CB, 2,5-CB, and 2,6-CB.

bData are means ± standard deviation (n = 3).

Table 5: Change in the number of chlorines per biphenyl over thefive month incubation period for the bioaugmentation experimentswith DF-1.

Location

Chlorines per biphenyl

Day 0 Day 145

No DF1 With DF1

Meb10 6.34± 0.02 6.30± 0.19 5.94± 0.10

Meb20 6.42± 0.02 6.29± 0.09 6.30± 0.04

Meb2 6.43± 0.01 6.42± 0.01 6.42± 0.01

and PCB-202 that did not contain any double-flanked metaor para positioned chlorines. However, the latter congenercoeluted with PCB-156 (50%-50% split) that contains bothdouble-flanked meta and para positioned chlorines. In thesamples Meb20 and Meb2 containing higher concentrationsof total PCB, bioaugmentation with DF-1 did not show anysignificant effects (data not shown).

4. Discussion

Bacteria capable of aerobic PCB degradation are foundubiquitously in the soil environment and several havebeen isolated and identified that belong to genera such

0

5

10

15

20

25

30

35

40

45

Mono Di Tri Tetra Penta Hexa Hepta Octa Nona DecaPCB homolog

Day 0No DF1-day 145With DF1-day 145

Mol

%

Figure 3: Homolog distribution showing the effect of DF-1 in thebioaugmentation experiment compared to effect of the indigenouspopulation for Meb10.

as Pseudomonas, Burkholderia, Ralstonia, Achromobacter,Comamonas, Bacillus and Rhodococcus [25, 44]. Althoughanaerobic dechlorination activity has been detected in soilsby activity based methods at several Canadian military instal-lations such as Saglek, Labrador [45], Resolution Island,


Nunavut [30], and Fort Albany, Ontario [46], anaerobic PCBdechlorinating bacteria have not been identified previouslyfrom soils. This is the first study to combine microbialcommunity analysis and activity-based techniques on PCBimpacted soils. The presence of active dechlorinating bacteriais an essential first step of the sequential degradationprocess that requires reduction of extensively chlorinatedPCB congeners to less chlorinated congeners that can besubject to aerobic ring cleavage by microbial 2,3- and 3,4-dioxygenase activity and can subsequently be mineralized byother aerobic microorganisms.

In the drainage ditch soil examined in this study, thePCB contamination originated predominantly from A1260contamination [42]. The presence of tetra chlorinatedhomologs atypical of Aroclor 1260 combined with identi-fication of the putative PCB dechlorinating phylotypes anddetection of PCB dechlorination activity suggests that insitu dechlorination occurs at this site. The 13 identifiedphylotypes clustered within the dechlorinating Chloroflexigroup, either within the the Dehalococcoides clade [47–49]within the broader Chloroflexi clade, which includes otherconfirmed PCB dechlorinating bacteria such as Dehalobiumchlorocoercia DF-1 and strain o-17-group [50]. In two previ-ous reports on bacterial communities in PCB contaminatedsoil, bacteria closely related to Proteobacteria, the Holophage-Acidobacterium phylum, Actinobacteria, and Plantomycetalesand Cytophagales were identified [51, 52]. Interestingly, thedominant species included the genera Burkholderia andVariovorax together with Sphingomonas species, Rhodophilaglobiformis group members, and Acidobacterium capsulatumthat aerobically degrade a variety of organic pollutantsincluding PCBs. However, neither anaerobic dechlorinatingbacteria nor any phylotypes related to the dechlorinatingChloroflexi were identified.

Prior to the current study, PCB dechlorinating bacteriahave only been identified in sediments. Thus, detectionof putative anaerobic dechlorinating bacteria and reductivedechlorination activity in the examined soil samples indi-cates that microbial dehalorespiration of PCBs also occursin soil. Putative aerobic PCB degrading bacteria were alsodetected at several locations in the drainage ditch, whichsuggests that natural attenuation of PCBs could occur in soilsby sequential anaerobic dechlorination followed by aerobicdegradation. However, the large spatial heterogeneity of bothanaerobic and aerobic phylotypes associated with PCB trans-formation could be an obstacle to bioremediation. Similarheterogeneity observations have been reported in studies ofsoils contaminated with heavy metals [9] and polycyclic aro-matic hydrocarbons [11]. In the first study, it was found thatsamples collected within 1 cm distance could have a 10,000fold difference in metabolic potential and that metal con-centrations did not correspond with the metabolic potential[9]. In another report, microscale heterogeneities observedin PAH contaminated soil were attributed to the particle sizeof sand and silt fractions, which impacted the PAH concen-trations and availability [11]. This corresponds with obser-vations of the behavior between bacteria and clay mineralsin soils contaminated with PCBs [12]. Here, it was observedthat soil aggregates called “clay hutches” housed the bacteria

creating a microhabitat, which would limit the PCB availabil-ity and constitute the main carbon source in an otherwisecarbon limited environment. These observations of spatialheterogeneity on both a macro- and microscale are consistentwith the observations of the spatial variability of bacteriainvolved in PCB transformation in soil in the current study.

The numbers of putative dechlorinating bacteria were2 to 5 orders of magnitude lower in soils compared withsediments, where up to approximately 108 bacteria per gsediment have been reported [33]. Dechlorination rates werelower in the soil compared to sediment indicating that theoverall potential for dechlorination is reduced. The lack ofcorrelation between the number of dechlorinating bacteriaand the dechlorination rates in the soil samples might havebeen caused by localized environmental conditions. Forexample, most of the soil samples at the time of samplingwere moist but were not submerged in water as sedimentsamples. In a study of simulated dredged sediment that wasspiked with 300 ppm A1248, it was reported that the dechlo-rination activities were lowered when the water content wasreduced, which caused a lag in the dechlorination of A1248[53]. Degradation studies of other contaminants in soil suchas petroleum hydrocarbons also show that moisture contentin soil influences the degradation rate [54].

A possible consequence of the reduced moisture contentin the soil samples could be exposure to oxygen thatnegatively influences anaerobic dechlorination. In a studyof methanogenic granules capable of A1254 dechlorination,exposure to oxygen did not significantly influence thedechlorination activity [55]. In six months, 80% of theinitial concentration of A1254 had been dechlorinated insamples that had been exposed to oxygen for one week. Thereason for this rapid dechlorination could be the formationof biofilms in microniches that stay anaerobic despite theexposure to oxygen at the surface of the granule. Micronichesin soil can result in spatial and temporal heterogeneity forparameters such as oxygen, pH, redox, and nutrients aswell as PCBs [12, 54]. Thus, another factor affecting thespatial variation in dechlorination rates observed in thesoil samples could be the heterogeneous moisture contentaffecting formation of biofilms and anaerobic micronichesthat support communities of dechlorinating bacteria.

Successful bioaugmentation of weathered Aroclor 1242in soil with an enrichment culture from the Hudson Riverhas been observed previously [56]. In the report, meta-dechlorination was observed after 19 weeks of incubationresulting in a reduction of the average chlorine content by0.7 chlorines per biphenyl. Another report that evaluatedthe effect of sequential anaerobic-aerobic treatment ofA1260 contaminated soil from Saglek, Labrador, Canada,[45] initiated with a dechlorinating enrichment, showed adecrease in the average chlorine content of 1.2 chlorinesper biphenyl together with a significant homolog shift afterthree months of anaerobic bioaugmentation [45]. In thecurrent study, bioaugmentation of A1260 contaminated soilwith a pure culture of D. chlororcoercia has resulted inonly 0.4 chlorines per biphenyl after 20 weeks. As expected,dechlorination of doubly flanked meta- and para-positionedchlorines was observed, but at least one product was


observed that could not have been formed due to DF-1activity, which suggests that D. chlororcoercia might havestimulated dechlorination of single-flanked chlorines by theindigenous dechlorinating population, possibly as a resultof priming of the indigenous populations by intermedi-ates products. Stimulation of dechlorinating activity afterbioagumentation with D. chlororcoercia has been reportedpreviously [57]. The relatively low rate of dechlorination insoil form Meb10 and lack of enhanced dechlorination at thetwo other bioaugmented sites might have been caused byinhibition by other contaminants such as metals and/or otherorganic contaminants [58].

In this study, the spatial distribution of bacteria involvedin PCB transformation was investigated in Aroclor 1260contaminated soil containing heavy metals and organiccocontaminants. The results showed that anaerobic PCBdehalorespiring bacteria were present and active. Detectionof genes encoding biphenyl dioxygenase genes suggeststhat the potential exists for natural attenuation by sequen-tial anaerobic dehalorespiration and aerobic degradation.Although a recent study demonstrated the potential forin situ treatment of PCB impacted sediment by bioaug-mentation [57], the spatial heterogeneity inherent in soilin the current study had a profound effect on the abilityto transform PCBs as only one of the bioaugmentedsamples showed positive effect. Strategies for effective insitu bioremediation of PCB-impacted soils may requireadditional remedial treatments such as flooding the soilprior to bioaugmentation to provide homologous watercontent, addition of a carbon source to create an adequatelyreduced environment for dehalorespiration, and precipita-tion of potentially inhibitory heavy metals or concurrentor sequential bioremediation with biocatalysts specificallytarget potentially inhibitory organic co-pollutants. Theresults illustrate some of the challenges associated with thedevelopment of in situ bioremediation strategies for PCBimpacted soils.

Acknowledgments

This work was supported in part by the Office of NavalResearch, U.S. Department of Defense, Grant N000014-03-1-0035 to K. R. Sowers and Grant N000014-03-1-0034 toH.D.May; U.S. Department of Defense, Strategic Environ-mental Research, and Development Program Project Num-bers ER-1502 and ER-1492 to K. R. Sowers in addition toresearch funding to B. V. Kjellerup from Phillips Oral HealthCare for development of DHPLC analysis methods. Theauthors thank Installation Restoration Program ManagerJeffrey A. Henning for his support and help to obtain samplesand site information.

References

[1] R. D. Kimbrough, “Polychlorinated biphenyls (PCBs) andhuman health: An update,” Critical Reviews in Toxicology, vol.25, no. 2, pp. 133–163, 1995.

[2] A. Fischbein, J. Thornton, M. S. Wolff, J. Bernstein, and I. J.Selifoff, “Dermatological findings in capacitor manufacturing

workers exposed to dielectric fluids containing polychlori-nated biphenyls (PCBs),” Archives of Environmental Health,vol. 37, pp. 69–74, 1982.

[3] H. K. Ouw, G. R. Simpson, and D. S. Siyali, “Use and healtheffects of Aroclor 1242, a polychlorinated biphenyl, in anelectrical industry,” Archives of Environmental Health, vol. 31,no. 4, pp. 189–194, 1976.

[4] S. Safe, K. Connor, K. Ramamoorthy, K. Gaido, and S.Maness, “Human exposure to endocrine-active chemicals:Hazard assessment problems,” Regulatory Toxicology andPharmacology, vol. 26, pp. 52–58, 1997.

[5] S. T. Vater, S. F. Velazquez, and V. J. Cogliano, “A case study ofcancer data set combinations for PCBs,” Regulatory Toxicologyand Pharmacology, vol. 22, no. 1, pp. 2–10, 1995.

[6] S. De Flora, M. Bagnasco, and P. Zanacchi, “Genotoxic,carcinogenic, and teratogenic hazards in the marine envi-ronment, with special reference to the Mediterranean Sea,”Mutation Research, vol. 258, no. 3, pp. 285–320, 1991.

[7] K. Kannan, H. Nakata, R. Stafford, G. R. Masson, S. Tanabe,and J. P. Giesy, “Bioaccumulation and toxic potential ofextremely hydrophobic polychlorinated biphenyl congeners inbiota collected at a superfund site contaminated with Aroclor1268,” Environmental Science and Technology, vol. 32, no. 9, pp.1214–1221, 1998.

[8] C. Backe, I. T. Cousins, and P. Larsson, “PCB in soils andestimated soil-air exchange fluxes of selected PCB congenersin the south of Sweden,” Environmental Pollution, vol. 128, no.1-2, pp. 59–72, 2004.

[9] J. M. Becker, T. Parkin, C. H. Nakatsu, J. D. Wilbur, andA. Konopka, “Bacterial activity, community structure, andcentimeter-scale spatial heterogeneity in contaminated soil,”Microbial Ecology, vol. 51, no. 2, pp. 220–231, 2006.

[10] W. A. Lead, E. Steinnes, J. R. Bacon, and K. C. Jones,“Polychlorinated biphenyls in UK and Norwegian soils: spatialand temporal trends,” Science of the Total Environment, vol.193, no. 3, pp. 229–236, 1997.

[11] W. Wilcke, W. Zech, and J. Kobza, “PAH-pools in soils alonga PAH-deposition gradient,” Environmental Pollution, vol. 92,no. 3, pp. 307–313, 1996.

[12] H. Lunsdorf, R. W. Erb, W. R. Abraham, and K. N. Timmis,“‘Clay hutches’: a novel interaction between bacteria and clayminerals,” Environmental Microbiology, vol. 2, pp. 161–168,2000.

[13] B. Ramakrishnan, T. Lueders, R. Conrad, and M. Friedrich,“Effect of soil aggregate size on methanogenesis and archaealcommunity structure in anoxic rice field soil,” FEMS Microbi-ology Ecology, vol. 32, no. 3, pp. 261–270, 2000.

[14] J. L. M. Rodrigues, C. A. Kachel, M. R. Aiello et al., “Degrada-tion of aroclor 1242 dechlorination products in sediments byBurkholderia xenovorans LB400(ohb) and Rhodococcus sp.strain RHA1(fcb),” Applied and Environmental Microbiology,vol. 72, no. 4, pp. 2476–2482, 2006.

[15] A. C. Alder, M. M. Haggblom, S. R. Oppenheimer, andL. Y. Young, “Reductive dechlorination of polychlorinatedbiphenyls in anaerobic sediments,” Environmental Science andTechnology, vol. 27, no. 3, pp. 530–538, 1993.

[16] P. B. Hatzinger and M. Alexander, “Effect of aging ofchemicals in soil on their biodegradability and extractability,”Environmental Science and Technology, vol. 29, no. 2, pp. 537–545, 1995.

[17] J. A. Field and R. Sierra-Alvarez, “Microbial transformationand degradation of polychlorinated biphenyls,” EnvironmentalPollution, vol. 155, no. 1, pp. 1–12, 2008.


[18] D. H. Pieper and M. Seeger, “Bacterial metabolism of poly-chlorinated biphenyls,” Journal of Molecular Microbiology andBiotechnology, vol. 15, no. 2-3, pp. 121–138, 2008.

[19] D. E. Fennell, I. Nijenhuis, S. F. Wilson, S. H. Zinder,and M. M. Haggblom, “Dehalococcoides ethenogenes Strain195 Reductively Dechlorinates Diverse Chlorinated AromaticPollutants,” Environmental Science and Technology, vol. 38, no.7, pp. 2075–2081, 2004.

[20] L. Adrian, V. Dudkova, K. Demnerova, and D. L. Bedard,“Dehalococcoides sp. strain CBDB1 extensively dechlorinatesthe commercial polychlorinated biphenyl mixture aroclor1260,” Applied and Environmental Microbiology, vol. 75, pp.4516–4524, 2009.

[21] L. A. Cutter, J. E. M. Watts, K. R. Sowers, and H. D. May,“Identification of a microorganism that links its growthto the reductive dechlorination of 2,3,5,6-chlorobiphenyl,”Environmental Microbiology, vol. 3, no. 11, pp. 699–709, 2001.

[22] S. K. Fagervold, H. D. May, and K. R. Sowers, “Microbialreductive dechlorination of Aroclor 1260 in Baltimore Harborsediment microcosms is catalyzed by three phylotypes withinthe Phylum chloroflexi,” Applied and Environmental Microbi-ology, vol. 73, no. 9, pp. 3009–3018, 2007.

[23] P. S. G. Chain, V. J. Denef, K. T. Konstantinidis et al.,“Burkholderia xenovorans LB400 harbors a multi-replicon,9.73-Mbp genome shaped for versatility,” Proceedings of theNational Academy of Sciences of the United States of America,vol. 103, no. 42, pp. 15280–15287, 2006.

[24] E. Masai, A. Yamada, J. M. Healy et al., “Characterizationof biphenyl catabolic genes of gram-positive polychlorinatedbiphenyl degrader Rhodococcus sp. strain RHA1,” Applied andEnvironmental Microbiology, vol. 61, no. 6, pp. 2079–2085,1995.

[25] K. Furukawa, “Molecular genetics and evolutionary relation-ship of PCB-degrading bacteria,” Biodegradation, vol. 5, no. 3-4, pp. 289–300, 1994.

[26] M. J. Hoostal, G. S. Bullerjahn, and R. M. L. McKay, “Molecu-lar assessment of the potential for in situ bioremediation ofPCBs from aquatic sediments,” Hydrobiologia, vol. 469, pp.59–65, 2002.

[27] R. W. Erb and I. Wagner-Dobler, “Detection of polychlo-rinated biphenyl degradation genes in polluted sedimentsby direct DNA extraction and polymerase chain reaction,”Applied and Environmental Microbiology, vol. 59, no. 12, pp.4065–4073, 1993.

[28] A. C. Singer, E. S. Gilbert, E. Luepromchai, and D. E. Crowley,“Bioremediation of polychlorinated biphenyl-contaminatedsoil using carvone and surfactant-grown bacteria,” AppliedMicrobiology and Biotechnology, vol. 54, no. 6, pp. 838–843,2000.

[29] S. Di Toro, G. Zanaroli, and F. Fava, “Intensification ofthe aerobic bioremediation of an actual site oil historicallycontaminated by polychlorinated biphenyls (PCBs) throughbioaugmentation with a non acclimated, complex source ofmicroorganisms,” Microbial Cell Factories, vol. 5, article no. 11,2006.

[30] B. Kuipers, W. R. Cullen, and W. W. Mohn, “Reductivedechlorination of weathered Aroclor 1260 during anaerobicbiotreatment of arctic soils,” Canadian Journal of Microbiology,vol. 49, no. 1, pp. 9–14, 2003.

[31] Q. Wu, K. R. Sowers, and H. D. May, “Establishment of a poly-chlorinated biphenyl-dechlorinating microbial consortium,specific for doubly flanked chlorines, in a defined, sediment-free medium,” Applied and Environmental Microbiology, vol.66, no. 1, pp. 49–53, 2000.

[32] U. Ghosh, A. S. Weber, J. N. Jensen, and J. R. Smith,“Relationship between PCB desorption equilibrium, kinetics,and availability during land biotreatment,” EnvironmentalScience and Technology, vol. 34, no. 12, pp. 2542–2548, 2000.

[33] B. V. Kjellerup, X. Sun, U. Ghosh, H. D. May, and K. R. Sowers,“Site-specific microbial communities in three PCB-impactedsediments are associated with different in situ dechlorinatingactivities,” Environmental Microbiology, vol. 10, no. 5, pp.1296–1309, 2008.

[34] M. Berkaw, L. Cutter, K. R. Sowers, and H. D. May, “Site-dependent ortho-, meta-, and para-dechlorination of PCBsby anaerobic estuarine and marine sediments enrichments,”Bergen, Norway, 1996.

[35] T. R. Pulliam Holoman, M. A. Elberson, L. A. Cutter, H. D.May, and K. R. Sowers, “Characterization of a defined 2,3,5,6-tetrachlorobiphenyl- orthodechlorinating microbial commu-nity by comparative sequence analysis of genes coding for 16SrRNA,” Applied and Environmental Microbiology, vol. 64, no.9, pp. 3359–3367, 1998.

[36] G. Muyzer, E. C. De Waal, and A. G. Uitterlinden, “Profilingof complex microbial populations by denaturing gradientgel electrophoresis analysis of polymerase chain reaction-amplified genes coding for 16S rRNA,” Applied and Environ-mental Microbiology, vol. 59, no. 3, pp. 695–700, 1993.

[37] D. J. Lane, B. Pace, G. J. Olsen, D. A. Stahl, M. L. Sogin, andN. R. Pace, “Rapid determination of 16S ribosomal sequencesfor phylogenetic analysis,” Proceedings of the National Academyof Sciences of the United States of America, vol. 82, no. 20, pp.6955–6959, 1985.

[38] R. Seshadri, L. Adrian, D. E. Fouts et al., “Genome sequence ofthe PCE-dechlorinating bacterium Dehalococcoides etheno-genes,” Science, vol. 307, no. 5706, pp. 105–108, 2005.

[39] M. Kube, A. Beck, S. H. Zinder, H. Kuhl, R. Reinhardt, andL. Adrian, “Genome sequence of the chlorinated compound-respiring bacterium Dehalococcoides species strain CBDB1,”Nature Biotechnology, vol. 23, no. 10, pp. 1269–1273, 2005.

[40] O. Faroon, L. Samuel Keith, C. Smith-Simon, and C. DeRosa, Polychlorinated Biphenyls: Human Health Aspects, WorldHealth Organization, Geneva, Switzerland, 2003.

[41] M. Berkaw, K. R. Sowers, and H. D. May, “Anaerobic orthodechlorination of polychlorinated biphenyls by estuarine sed-iments from Baltimore Harbor,” Applied and EnvironmentalMicrobiology, vol. 62, no. 7, pp. 2534–2539, 1996.

[42] Tetra Tech NUS I, 2005 Soil Sampling Summary Report forSite 9 - Stormwater Drainage Ditch, Bioremediation Pilot Study,Naval Support Activity, Pittsburg, Pa, USA, 2005.

[43] W. Ludwig, O. Strunk, R. Westram et al., “ARB: a softwareenvironment for sequence data,” Nucleic Acids Research, vol.32, no. 4, pp. 1363–1371, 2004.

[44] D. H. Pieper, “Aerobic degradation of polychlorinatedbiphenyls,” Applied Microbiology and Biotechnology, vol. 67,no. 2, pp. 170–191, 2005.

[45] E. R. Master, V. W. M. Lai, B. Kuipers, W. R. Cullen, and W.W. Mohn, “Sequential anaerobic - Aerobic treatment of soilcontaminated with weathered aroclor 1260,” EnvironmentalScience and Technology, vol. 36, no. 1, pp. 100–103, 2002.

[46] L. J. S. Tsuji, B. C. Wainman, I. D. Martin, J. P. Weber, C.Sutherland, and E. Nieboer, “Abandoned Mid-Canada RadarLine sites in the Western James region of Northern Ontario,Canada: a source of organochlorines for First Nations people?”Science of the Total Environment, vol. 370, no. 2-3, pp. 452–466,2006.

[47] A. S. Waller, R. Krajmalnik-Brown, F. E. Loffler, and E.A. Edwards, “Multiple reductive-dehalogenase-homologous


genes are simultaneously transcribed during dechlorinationby dehalococcoides-containing cultures,” Applied and Envi-ronmental Microbiology, vol. 71, pp. 8257–8264, 2005.

[48] D. L. Bedard, J. J. Bailey, B. L. Reiss, and G. Van Slyke Jerzak,“Development and characterization of stable sediment-freeanaerobic bacterial enrichment cultures that dechlorinatearoclor 1260,” Applied and Environmental Microbiology, vol.72, no. 4, pp. 2460–2470, 2006.

[49] E. R. Hendrickson, J. A. Payne, R. M. Young et al., “Molec-ular analysis of Dehalococcoides 16S ribosomal DNA fromchloroethene-contaminated sites throughout North Americaand Europe,” Applied and Environmental Microbiology, vol. 68,no. 2, pp. 485–495, 2002.

[50] S. K. Fagervold, J. E. M. Watts, H. D. May, and K.R. Sowers, “Sequential reductive dechlorination of meta-chlorinated polychlorinated biphenyl congeners in sedimentmicrocosms by two different Chloroflexi phylotypes,” Appliedand Environmental Microbiology, vol. 71, no. 12, pp. 8085–8090, 2005.

[51] B. Nogales, E. R. B. Moore, E. Llobet-Brossa, R. Rossello-Mora, R. Amann, and K. N. Timmis, “Combined use of 16Sribosomal DNA and 16S rRNA to study the bacterial commu-nity of polychlorinated biphenyl-polluted soil,” Applied andEnvironmental Microbiology, vol. 67, no. 4, pp. 1874–1884,2001.

[52] B. Nogales, E. R. B. Moore, W. R. Abraham, and K. N.Timmis, “Identification of the metabolically active membersof a bacterial community in a polychlorinated biphenyl-polluted moorland soil,” Environmental Microbiology, vol. 1,no. 3, pp. 199–212, 1999.

[53] Y. C. Cho, O. S. Kwon, R. C. Sokol, C. M. Bethoney, and G.Y. Rhee, “Microbial PCB dechlorination in dredged sedimentsand the effect of moisture,” Chemosphere, vol. 43, no. 8, pp.1119–1126, 2001.

[54] C. Davis, T. Cort, D. Dai, T. H. Illangasekare, and J. Munakata-Marr, “Effects of heterogeneity and experimental scale on thebiodegradation of diesel,” Biodegradation, vol. 14, no. 6, pp.373–384, 2003.

[55] M. R. Natarajan, H. Wang, R. Hickey, and L. Bhatnagar, “Effectof oxygen and storage conditions on the metabolic activities ofpolychlorinated biphenyls dechlorinating microbial granules,”Applied Microbiology and Biotechnology, vol. 43, no. 4, pp. 733–738, 1995.

[56] K. T. Klasson, J. W. Barton, B. S. Evans, and M. E. Reeves,“Reductive microbial dechlorination of indigenous polychlo-rinated biphenyls in soil using a sediment-free inoculum,”Biotechnology Progress, vol. 12, no. 3, pp. 310–315, 1996.

[57] R. B. Payne, H. D. May, and K. R. Sowers, “Enhanced reductivedechlorination of polychlorinated biphenyl impacted sedi-ment by bioaugmentation with a dehalorespiring bacterium,”Environmental Science & Technology, vol. 45, pp. 8772–8779,2011.

[58] D. A. Abramowicz, M. J. Brennan, H. M. van Dort, and E.L. Gallagher, “Factors influencing the rate of polychlorinatedbiphenyl dechlorination in Hudson River sediments,” Environ-mental Science & Technology, vol. 27, pp. 1125–1131, 1993.


Research Article

A Critical Evaluation of Single Extractions from the SMT Programto Determine Trace Element Mobility in Sediments

Valerie Cappuyns1, 2

1 Center for Economics and Corporate Sustainability (CEDON), University College Brussels (HUB), Warmoesberg 26,1000 Brussels, Belgium

2 Department of Earth and Environmental Sciences, KULeuven, Celestijnenlaan 200E, 3001 Heverlee, Belgium

Correspondence should be addressed to Valerie Cappuyns, [email protected]

Received 30 November 2011; Revised 27 March 2012; Accepted 4 April 2012


Copyright © 2012 Valerie Cappuyns. This is an open access article distributed under the Creative Commons Attribution License,which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

Two commonly applied single extractions procedures, namely extractions with ammonium-EDTA and acetic acid, were evaluatedbased on the analysis of 72 samples from alluvial sediments. For most trace elements (Cu, Zn, Cd, Ni, As, and Pb), a significantlinear relationship could be established between their ammonium-EDTA or acetic acid extractable concentrations and their totalconcentrations, the organic carbon content, pH, and Fe , Al, and/or Ca content in the sediments. The scientific understandingof trace element partitioning in the complex soil-water system with these simple models is rather limited, but they offer theopportunity to use data from single extractions in a more comprehensive way. Despite the fact that these extractions cannotdirectly be related to the bioavailability of elements, they can provide input data for use in risk assessment models. Additionally,they also offer possibilities to perform a fast screening of the mobilizable pool of elements in soils and/or sediments.

1. Introduction

The contamination of soils and sediments is widespread andis a potential threat for the environment in the short and longterm. The impact of trace elements in soils and sedimentson the environment depends on their speciation, mobility,and bioavailability. Over the past decades, the term “heavymetals” has increasingly been used, without any consistencyto denote trace element contamination of environmentalmedia. An overview of the use of the term “heavy metals”in scientific dictionaries and relevant literature can be foundin Duffus [1]. Since “heavy metals” is a poor scientific termand many alternatives exit [2], we will use the term “traceelements” in the present study to refer to As, Cd, Cu, Cr, Ni,Pb, and Zn. Talking about trace metals would be incorrectbecause arsenic is actually a metalloid.

Before discussing the different methods for determina-tion of “trace element” availability in soils and/or sedimentsand before addressing the pros and cons of single andsequential extraction procedures, the difference between soilsand sediments will be clarified, as well as the terminologyused throughout this paper.

1.1. Soils versus Sediments. Soils and sediments are differ-ent matrixes from many viewpoints, especially under theenvironmental context. “Soil” can be defined as a “three-dimensional body with properties that reflect the impactof climate, vegetation, fauna, and topography on soilsparent material over a variable time span. Soils are stillin a process of change. As a result of “soil formation”or “pedogenesis,” soil profiles show signs of differentiationor alteration of the soil material [3].” “Sediment” can bedescribed as “material that is transported by water andsettles down from the water column [4]. In freshly depositedalluvial sediments, signs of differentiation or alteration of thematerial are sometimes not yet observable.” Nevertheless, insoil classification, specific designations are foreseen for thiskind of “material”: alluvial soils can often be classified asFluvisols, which “exhibit a stratified profile that reflects theirdepositional history or an irregular layering of humus andmineral sediments in which the content of organic carbondecreases with depth [5].” The qualifiers fluvic and spolicare used to indicate, respectively, the regular deposition offresh sediments or the deposition of dredged sediments ona soil.


Throughout this paper, the term “sediment” will be usedto indicate both the river sediments and the alluvial soilsthat consist of dredged-sediment derived soils and overbanksediments, regardless of their specific origin (e.g., overbankflooding, dredged-sediment derived soils, etc.), the degree ofalteration, or the catchment width.

1.2. Trace Element Mobility in Soils and Sediments: Exper-imental Approach. “Trace element mobility” is an opera-tionally defined term, which is determined by an approachused to determine the mobile, labile, or available metalspecies in soils and sediments. Although spectroscopic toolssuch as X-ray adsorption fine structure (XAFS) spectroscopycan give information on the coordination chemistry ofmetals (e.g., [6]), the quantification of the most mobilespecies is still difficult.

The composition of soil pore water is important from anenvironmental point of view because it gives an indicationof the “actual mobility” of trace elements and because theuptake of trace elements by plants occurs via the pore water.Moreover, pore water is also the carrier for elements to thegroundwater. Leaching is the process by which inorganicor organic contaminants in the pore water are moved todeeper soil layers or to the groundwater by infiltratingwater. However, the pore water composition only gives amomentary picture of trace element mobility since porewater composition can change over time. To assess traceelement mobility in the long term (referred to as “poten-tial mobility,” including physicochemically and biologicallyavailable metal pools) and under changing environmentalconditions a variety of leaching and extractions tests are used.According to Peijnenburg et al. [7], three approaches can bedistinguished to quantify physicochemically and biologicallyavailable metal pools in the soil: (1) direct measurement ormodelling of metal activities, (2) assessment of operationallydefined element fractions by means of single and sequentialextractions, and (3) application of semipermeable devices,such as ion exchange resins/membranes and toxicity testswith membrane devices. For example, the in situ techniqueof diffusive gradients in thin films (DGT) is used formeasuring effective soil solution concentrations and theadditional element concentration supplied from the solidphase.

The direct measurement of metal activities in pore wateris rather complex, and there is not always an agreementbetween measured and model concentrations of free metalion activities [8]. Several investigations have also beenperformed to compare the results of diffuse gradients in thinfilms (DGT) with single and sequential extraction methods.Roulier et al. [9] compared the mobilisation of Co, Cd, andPb in sediments using DGT and sequential extractions andfound a correlation between the masses of metals trappedin DGT resins and the metals extracted during sequen-tial extractions. Other authors concluded that the DGT-methodology did not have an additional value in predictingbioavailability of zinc [10] or uranium [11] in terrestrialecosystems as compared to conventional extraction methods.Nevertheless, the DGT measure of trace elements providesa promising indicator of their toxicity [12], but interacting

effects should first be clarified before DGT can be used inroutinely risk assessment of soils and sediments.

The present work will focus on the second approach,with emphasis on two commonly applied single extractionsprocedures, which will be critically evaluated.

1.3. Single and Sequential Extractions. Single and sequen-tial extractions provide semiquantitative information onelement distribution between operationally defined geo-chemical fractions. Therefore, the fractions obtained fromsingle and sequential extraction do not necessarily reflecttrue chemical speciation. The different extractions are oftenintended to simulate processes in nature, such as acidifica-tion or oxidation. However, the physicochemical conditionsin single and sequential extraction experiments (strongreagents and rapid reactions) often differ from natural con-ditions (weak reagents and slow reactions) [13]. Althoughleaching techniques such as column leaching and pHstat

leaching tests are probably more realistic to field conditions,single sequential extractions can give an indication of the“pools” or “sinks” of trace elements that are potentiallyavailable under changing environmental conditions.

From a practical point of view, the main drawbacks ofsequential extraction procedures are that they are ratherlaborious and time-consuming. Moreover, not all the steps ina given procedure are equally important in soil or sedimentsamples with different composition. Sometimes, one is notinterested in the association of a metal with different phasesin soil but wants to estimate the environmental risks of traceelements, for example, the availability to plants. In recentyears, attempts to improve single and sequential extractionstowards higher selectivity and higher operational efficiencyhas been achieved. With respect to this more problem-orientated approach, Maiz et al. [14] proposed a shortextraction scheme only considering mobile (“exchangeable”:CaCl2), mobilizable (DTPA), and residual forms. GomezAriza et al. [15] developed an improved extraction schemefor heavily polluted and iron-oxide-rich sediments, usingrepetitive extractions with NH2OH·HCl 0.4 mol/L.

The most common problems with sequential extractionsare the nonselectivity of reagents and readsorption phenom-ena [16, 17]. Besides the measurement of elements in theextracts, the analysis of the solid phase (X-ray diffraction,energy dispersive spectroscopy, and microprobe analysis),after extraction with a reagent, can give information on theselectivity of the reagent and the completeness of reaction[18–24]. The extraction of model solid phases can also yieldinformation of the selectivity and efficiency of reagents inthe different steps of a sequential extraction scheme (e.g.,[15, 25, 26]). La Force and Fendorf [23, 24] concluded thatsequential extractions should not be universally applied to allsoils but need to be evaluated on a site basis for a given soil.Therefore, optimization of a given extraction (concentrationof reagents, sequence, and reaction time) is required. Onthe other hand, standardization of these procedures is theonly way to achieve comparability when using sequentialextractions [27]. In the past decade, much effort has beenmade to evaluate sequential extraction schemes. The bestexample is the BCR sequential extraction scheme, a simple


Table 1: Examples of different types of extractions with EDTA salts applied to soils and sediments.

Extractant Applied to t L/S [conc] Metals analyzed Reference

(NH4)4-EDTA Sediments 1 h 10 mmol/L [38]

Na2H2-EDTA Soils 2 h 5 10 mmol/L Cd, Zn, Pb, Cu, and Ni [37]

Na2H2-EDTANa4-EDTA

Calcareous soils 22 h 10 250 mmol/L Cd, Pb, Zn, and As [39]

H4-EDTA Urban soils 24 h 5 68.4, 13.7 and 27.4 mmol/L Cd, Pb, Zn, and Cu [41]

Na2Ca-EDTA Calcareous soils Pb [42]

3-stage procedure that was thoroughly tested by interlabora-tory trials. The original procedure [28] consisted of 3 extrac-tions that separated “acid extractable” (CH3COOH 0.11 M),“reducible” (NH2OH·HCl 0.1 M, pH 2), and “oxidizable”(H2O2 15%) fractions. During the certification of referencematerials [29, 30], the reducing extraction (NH2OH·HCl)in the BCR sequential extraction scheme was found tosuffer from a lack of reproducibility. After testing differentreaction conditions (concentration of the reagent, pH), theNH2OH·HCl concentration was changed to 0.5 M and thepH of the reagent was adjusted to 1.5 by the addition of afixed volume of HNO3.

Finally, different extraction procedures, applied on thesame sample, are often compared to select the procedurethat is most suited for the soil or sediment of concern (e.g.,[15]). However, the direct comparison between methodsis difficult to carry out, especially when different reagentsare applied to extract a specific phase or when reagentswith different concentrations are used in the methods to becompared.

1.4. Single Extractions Applied in the Present Study. Singleextractions represent a relatively fast, cheap, and simple wayto assess trace element mobility in contaminated soils andsediments. Depending on the objectives of the extraction,water, diluted salt solutions, or stronger reagents such asEDTA are used. Single extractions are also widely appliedin soil science for the quantification of the amount ofFe and Al oxides in soils. Gupta et al. [31] presented arisk assessment and risk management guideline concept tohandle contaminated soils or sites. Mobile and mobilizablemetal fractions were introduced, which can be separatedby means of single extractions. These fractions are oper-ationally defined by the method used to separate them.The mobile fraction is equivalent to the “actually available”metal fraction, while the mobilizable fraction is relatedwith the potential availability of trace elements in soils andsediments.

1.4.1. EDTA. EDTA extractions are often used to estimatethe potentially available pool (i.e., the pool that can delivermetals from the solid phase of the soil to the soil solutionin a relatively short time period). EDTA exhibits a strongcapacity to complex metals. EDTA was shown to dissolvecarbonates, thereby mobilizing occluded elements [32].Borggaard [33] showed that EDTA extracts amorphous Feoxides, but this dissolution is very slow in the presenceof other metal-chelate complexes [34]. It is also able to

form organometal complexes, which compete with organicmatter in soil. Several authors mention that an extractionwith EDTA provides results similar to the sum of all of theextractable metals in a sequential extraction scheme (totalcontent minus the load associated with the residual phase)[35].

Different types of EDTA salts are used to extract soilsand sediments, in different concentrations and at differentsolid/liquid ratios. Additionally, EDTA can be applied tosoils and sediments by percolation in column leaching tests(e.g., [36, 37]) or in batch extractions. Some examples of thedifferent operational conditions of batch EDTA extractionsare given in Table 1. Sodium-EDTA and ammonium-EDTAare the most frequently used EDTA salts, whereas EDTA as afree acid is rarely used to extract soils and sediments.

Zou et al. [38] investigated the influence of differentEDTA salts (EDTA free acid, Na-EDTA, and NH4-EDTA)at different concentrations and pH values. The extractionefficiency of EDTA decreased with increasing pH in the pHrange 2–10, and consecutive extractions with diluted EDTAsolutions are more effective than a single extraction witha concentrated solution [38]. The tetrasodium salt (Na4-EDTA) is less effective for heavy metal removal compared tothe disodium salt (Na2H2-EDTA) [39]. According to Finzgarand Lestan [40], multiple dosages of EDTA were substantiallymore effective for leaching Pb from contaminated soils thanusing one large single dose. Similarly, two extractions withEDTA (5 mmol EDTA/kg) removed more Cu than a singledose of 10 mmol/kg EDTA.

Additionally, besides the type of EDTA salt, the pH,and the liquid/solid ratio, time is an important variablethat influences the results of EDTA extractions. Bermondand Ghestem [43] used a kinetic fractionation method tomonitor the EDTA extraction of soil elements versus time.They assumed the existence of two sorts of metallic cations,that is, labile cations-that were quickly extracted, and theslowly or moderately labile cations that were less quicklyreleased. The reactions were considered as pseudo-first-orderreactions since the EDTA reagent was in excess and describedby a first-order equation [44].

The effectiveness of EDTA salts to extract trace elementsfrom soils and sediments is very variable, not only dependingon extraction conditions but also on the composition of soilsand sediments and on the speciation of trace elements.

Because of the relatively low cost of EDTA, soil washingwith EDTA, both as an in situ and ex situ remediationtechnique, has been evaluated by many researchers.


EDTA can effectively remove Cd, Cu, Pb, and Zn fromsoils with removal efficiencies ranging between 65 and 86%[45]. However, due to possible adverse health and environ-mental effects, the use of EDTA is currently under scrutiny.Kalf et al. [46] determined the maximum permissibleconcentration (MPC) and negligible concentration (NC) forEDTA in water, based on the EU risk assessment report forthis compound. The maximum permissible concentration(MPC) for EDTA in water is 2.2 mg/L, and the negligibleconcentration (NC) is 0.022 mg/L. Calculation of MPCsfor sediment or soil was not possible due to the complexspeciation of EDTA in soils and sediments.

Since the toxic wastewaters that are generated during soilwashing with EDTA cannot be treated using conventionalmethods such as filtration, flocculation, and precipitation[47], several methods such as electrochemical advancedoxidation [48] or resin trapping techniques [41] have beentested to recuperate the washing solutions or to clean thewastewaters containing EDTA.

1.4.2. Acetic Acid. Extractions with acetic acid or with acetatesalts are often carried out as one of the first steps in asequential extractions scheme. A 1 mol/L sodium acetatesolution, acidified to pH 5, and 0.11 mol/L acetic acid are themost widely used reagents to determine “acid extractable”metal concentrations in sequential extractions schemes:sodium acetate is used to determine the “carbonate fraction”in the Tessier sequential extraction procedure [49], whereasthe first step of the BCR sequential extraction procedure [50]consists of an extraction with 0.11 mol/L acetic acid.

One of the major parameters that influence the selec-tivity of reagents during single and sequential extractionsis the pH of the extracts. 0.11 mol/L acetic acid (step 1in the BCR sequential extraction scheme) is supposed torelease “exchangeable” elements and to dissolve some poorlycrystalline hydroxy- and carbonate-metal phases [32]. It isof major concern that this extraction is as complete aspossible since pH and changes in pH during extractions arekey parameters that determine the potential redistributionof trace elements during (sequential) extractions [43]. Theextraction with acetic acid 0.43 mol/L allows an estimationof the metal fraction remobilized after acidification of the soilto the pH of the extracting agent [51].

1.4.3. Single Extractions of the SM&T Program. The Euro-pean Standards, Measurements and Testing program (SMT),formerly BCR (European Community Bureau of Referencein Brussels) supports standardization. Between 1996 and2000, BCR has sponsored studies of soil extraction proce-dures for trace element speciation conducted by a WorkingGroup of some 35 European laboratories. These studiesincluded the evaluation of single extracting agents for traceelement speciation. Within the framework of harmonizationof leaching procedures for risk assessment of trace elementsin soils, the SM&T performed an extensive collaborativestudy, which resulted in the selection of ammonium-EDTA0.05 mol/L and CH3COOH 0.43 mol/L (extract the “mobiliz-able fractions,” indicating the “potential availability”) as soilextracting agents [50].

In the present study, “potentially mobile” fractions oftrace elements in a set of 72 sediment samples with differentdegree of contamination, major element composition, grainsize, organic carbon content, and pH, were separated bymeans of single extractions that were recommended by theSM&T of the European Commission. The results of bothextractions were compared, and the relationships betweenthe extractability of trace elements and the compositionof the sediments samples (degree of contamination, majorelement composition, grain size, organic carbon content, andpH) were investigated and quantified when possible. Finally,different possible uses and interpretations of the acetic acidand ammonium-EDTA extractable trace element fraction arecritically discussed.

2. Methodology

2.1. Sampling and Sample Pretreatment. The samples ana-lyzed in this study consist of river sediments, alluvial soilsamples (land disposed dredged sediments and overbanksediments) from 10 different locations in Flanders (NorthernBelgium). Throughout this paper, the term “sediment” willbe used to indicate both the river sediments and the alluvialsoils that consist of dredged-sediment derived soils andoverbank sediments. The river sediments (3 samples) weresampled with a Van Veen Grab, taking the uppermost 5 cmof the sediment surface. For the alluvial soil samples, profilepits were dug until the depth of the water table and 3 to5 samples were taken, depending on visual differences incolor and texture. During sampling, precaution was taken tominimize metal contamination from the grab or the spade,for example, the outer part of the sediment sample wasremoved and only the inner part was further processed.In total, 72 samples were collected in plastic bags andtransported to the laboratory, where they were air-dried. Forthe physicochemical analysis, part of the sample was gentlydisaggregated in a porcelain mortar and sieved (<2 mm).

2.2. Physicochemical Sample Characteristics. pH(H2O) wasmeasured in a soil/water suspension (1/2.5 kg/L) (pH Hamil-ton single-pore electrode). Organic carbon was determinedaccording to the Walkley and Black method [52], andeffective cation exchange capacity (CEC) was analysed forapplying the “silver thiourea method” [53, 54]. Grain sizecomposition was determined by means of laser diffrac-tion spectrophotometry (Malvern Mastersizer S long bed,Malvern, Worcestershire, UK). For practical reasons, CECand grain size distribution analysis were only performed on34 samples. Total element concentrations (Al, As, Cd, Co, Cr,Cu, Ni, Pb, Zn, Fe, Mg, Mn, K, and Ca) were determinedafter dissolution of the samples with a mixture of threeconcentrated acids (4 mL HClconc, 2 mL HNO3conc, and 2 mLHFconc).

2.3. Reagents. All reagents used were of analytical gradeSuprapur quality (acetic acid from Riedel-de Haen, NH3 andEDTA from Merck).

The 0.05 mol/L ammonium-EDTA extracting solutionwas prepared as an ammonium salt solution by adding in


a fume cupboard 146.12± 0.05 g of EDTA free acid to 800±20 mL distilled water and by partially dissolving by stirringin 130 ± 5 mL of concentrated ammonia solution until allthe EDTA was dissolved. The obtained solution was filtered(2.0 μm) into a 10-litre polyethylene container and dilutedwith water to 9.0± 0.5 L. The pH was adjusted to 7.00± 0.05by addition of a few drops of hydrochloric acid. Finally, thesolution was diluted with distilled water to 10 ± 0.1 L, wellmixed, and then stored in stoppered polyethylene container.

The 0.43 mol/L acetic acid extracting solution was pre-pared by adding 250 ± 2 mL of redistilled glacial acetic acidin a fume cupboard to about 5 litres of distilled water in a10 L polyethylene container. The solution was diluted withdistilled water to 10 L volume, well mixed, and then stored ina stoppered polyethylene container.

Standard solutions for FAAS were made by serial dilutionof 1000 μg/L standard solutions (Merck) of the appropriateelements. Standard series for ICP-MS were made, startingfrom the “10 ppm multielement calibration standard 2A in5% HNO3” from Hewlett Packard.

2.4. Single Extractions. For the ammonium-EDTA and aceticacid extraction, the protocol of the SM&T program [27] wasfollowed. 20 mL of a 0.05 mol/L ammonium-EDTA solutionwas added to 2 g of dry sediment. The suspension was shakenfor 1 h in a reciprocal shaker, centrifuged (3500 rpm, 10 min),decanted off, and filtered (0.45 μm). 40 mL of a 0.43 mol/LCH3COOH solution was added to 1 g of dry sediment.The suspension was shaken for 16 h in a reciprocal shaker,centrifuged (3500 rpm, 10 min), decanted, and then filtered(0.45 μm). After measuring the pH, the CH3COOH extractswere acidified with concentrated HNO3 to bring the pH <2. The EDTA extracts were not acidified prior to analysisto prevent precipitation of EDTA salts at very low pH. AReference material (CRM 483) certified for its ammonium-EDTA and acetic acid extractable content of Cd, Cr, Cu,Ni, Pb, and Zn was also included in quadruplate. Blankextractions (i.e., without soil or sediment) were carried outfor each set of analysis, using the same reagents as describedabove.

2.5. Analysis of the Leachates. The solutions were analysedby flame atomic adsorption spectrometry (FAAS) (VarianTechtron AA6) for Ca, Fe, K, and Al. For As, Cd, Cr, Cu, Mn,Ni, Pb, and Zn a multielement analysis by ICP-MS (HP 4500series) was carried out. All samples were diluted with HNO3

1 mol/L (ultra pure) to maintain a comparable matrix for allsamples.

For measurement of the single extraction leachates,standard solution series were made in such away that theycontained the same proportion of the respective extractionsolution and background. All glassware was thoroughlycleaned with HNO3 0.15 mol/L. Reagent blanks were deter-mined for each new batch of reagent. Detection limitswere calculated according to the procedure in the HP4500 Application Handbook. Except for Ca (10 μg/L) andfor Fe (3 μg/L), the detection limit was generally below1 μg/L. An indium (In) internal standard was applied toboth samples and standards to minimise the influence of

nonspectroscopic interferences such as signal suppression orenhancement when measuring samples with a high matrixconcentration. The spectroscopic interference of ArCl, whichhas the same m/z as As (75), was corrected according tothe recommendations of the US Environmental ProtectionAgency (EPA) (method 200.8, [55]). The pH of the extractswas measured with a pH Hamilton single-pore electrode.

2.6. Statistical Analysis. Statistical analysis was performedwith the software package SPSS 18.0 for Windows. Descrip-tive statistics (average, median, minimum, maximum, andstandard deviation) were calculated for each variable. Thenormal distribution of the variables was checked by meansof the Kolmogorov-Smirnov (K-S) test, and correlationsbetween variables were tested by calculating two-tailedPearson’s correlation coefficients. An α-value of 0.01 wasadopted at the critical level for all statistical testing, givinga 99% confidence level. One-way ANOVA was performedto investigate whether there was a difference between bothreagents concerning the ability to extract elements from thesediments.

Finally, stepwise multiple linear regression was per-formed to deduce possible causal relationships betweenthe variables. Attention was mainly paid to the possi-bility of predicting trace element concentrations in theammonium-EDTA and acetic acid extracts based on traceelement content, major element composition, pH, andorganic matter content. Different assumptions of the linearregression (normality of the de residues, autocorrelation,quasi-multicollinearity (QMC), and heteroscedasticity) weretested.

For the statistical analysis, the recommendations ofWebster [56, 57] (reporting mean values with standarderrors, performing linear regressions, etc.) were taken intoaccount.

3. Results and Discussion

3.1. Certified Reference Materials. Certified reference mate-rials, which are also called standard reference materials,are used for the verification of the accuracy of analyticalprocedures following strict extraction protocols [58]. In thepresent work, single extractions (with ammonium-EDTA0.05 mol/L, acetic acid 0.43 mol/L) were performed on twocertified reference materials that were developed within theframework of the SM&T program, namely, BCR 701 andCRM483.

Total concentrations of Cd, Cr, Cu, Ni, Pb, and Znwere determined in reference material BCR-701 [59] bydissolution with 3 concentrated acids (HF, HCl, and HNO3)(Table 2). Although the extraction procedure is differentfrom the aqua regia extraction, the own values were within1 standard deviation of the indicative values, except for Cu.

CRM 483 is a sewage-sludge amended soil that was col-lected by multiple sampling to a depth of 10 cm [60]. Withinthe framework of harmonization of leaching procedures forrisk assessment of trace elements in soils (SMT program), thereference material CRM 483 was certified for its ammonium-EDTA and acetic acid extractable contents [28].


Table 2: Comparison between aqua regia extractable (indicativevalues) and “3-acid” extractable (this work) concentrations (mg/kg)of Cd, Cr, Cu, Ni, Pb, and Zn in sample BCR 701. Mean ± standarddeviation of 3 replicates.

Indicative values This work

Cd 11.7± 1.0 10.9± 0.2

Cr 272± 20 284± 5

Cu 275± 13 242± 2

Ni 103± 4 99± 7

Pb 143± 6 141± 3

Zn 454± 19 465± 4

Table 3: Comparison of the results of the ammonium-EDTA andacetic acid extractions of CRM 483 with certified and indicativevalues [31]. The number of replicates is given between brackets.Concentrations in mg/kg.

CRM483 Certified values [31]

Ammonium-EDTA Acetic acid

Cd 20.4± 1.3 18.3± 0.6

Cr 28.6± 2.6 18.7± 1.0

Cu 215± 11 33.5± 1.6

Ni 28.7± 1.7 25.8± 1.0

Pb 229± 8 3.1± 0.25

Zn 612± 19 620± 24

CRM483 This work

Ammonium-EDTA (6) Acetic acid (6)

Cd 20.6± 0.6 15.2± 1.48

Cr 26.4± 3.4 13.3± 1.95

Cu 207± 4 28.1± 0.46

Ni 26.5± 1.78 21.5± 0.77

Pb 202± 27 1.7± 0.21

Zn 607± 4 556± 23

A good agreement was obtained between the ownand indicative/certified values for the ammonium-EDTAextractions (Table 3). Experimental values always differed byless than 1 standard deviation from the certified values.

The results of the acetic acid extraction, however, weresystematically lower than the certified values. This canbe related to the fact that a reciprocal shaker was usedinstead of an end-over-end shaker, as recommended bythe SM&T procedure. The shaker speed and the use of anend-over-end shaker (instead of a reciprocal shaker) areconsidered to be important parameters since they representfactors that condition the maintenance of the samples insuspension during extraction. During the comparison oftrace element extractability in sample CRM 483, the resultsobtained from the reciprocal shaker were systematically toolow [31]. Therefore, the centrifuge bottles were placed in theshaking device with an inclination of approximately 45◦. Thismodification allowed a better suspension of the sedimentsduring extraction. For the ammonium-EDTA extraction, for

which 2 g of material was suspended in 20 mL of solutionin a 50 mL centrifuge tube, the samples were maintainedin suspension during extraction. However, during the aceticacid extraction, 1 g of material was suspended in 40 mLof solution in a 50 mL centrifuge tube, resulting in a lessefficient suspension of the material because of the higheramount of liquid in the centrifuge tube. This may explainthe deviating results of the acetic acid extractions.

3.2. General Sample Characteristics. The 72 samples repre-sent a variety of sediments with different contents of majorelements, trace elements, organic carbon, and pH. The pH ofthe samples investigated in this study was in the range 5.3–8.1(Table 4). The grain size distribution of the samples variedfrom clayey to sandy, and cation exchange capacity (CEC)was between 7 and 38 cmol/kg (not in Table 4). A significantpositive linear correlation (R = 0.760) was found betweenCEC, on the one hand, and the organic carbon and Alcontent (which can be used as a proxy for the clay content),on the other hand. Since CEC was not determined for all thesamples, it was not included in the further analysis of thedataset. All the samples were polluted with at least one ofthe following trace elements: Zn, Cd, Pb, Cu, As, and/or Ni.Significant correlations were found (0.01 significance level)between most trace elements (Cr, Cu, Zn, Ni, Cd, Pb, and Co,R = 0.805–0.382). Ca and pH were also positively correlated(R = 0.789), as well as Fe and the organic carbon content (R =0.456) and Ca and pH (R = 0.860).

With respect to the extraction with ammonium-EDTA,Zn, Cd, Cu, Ni, and As correlated positively with each other,whereas the correlation between pH and EDTA-extractableAl and Fe was negative. Finally, for the acetic acid extraction,a positive correlation was found between the extractedconcentrations of Cr, Cu, Zn, Co, and Mn.

3.3. Extraction Efficiency. Single extractions with acetic acidand ammonium-EDTA provide some information on theinfluence of acidification and complexation on trace elementmobility. In general, Zn, Cd, and Ni were sensitive to bothacetic acid and ammonium-EDTA, whereas Pb and Cu weremore sensitive to an extraction with ammonium-EDTA. Thehighest average extraction efficiency with acetic acid wasobtained for Cd and Zn, while Cu was most effectivelyextracted with ammonium-EDTA.

The extraction efficiency of both extracting agents wascompared in order to assess their ability to release metalsfrom the “weakly bound” element pool of the sediments.

The results from the one-way ANOVA (Table 6) showedthat there were significant differences between both extract-ing agents for Cu, Pb, and Co and for most of the majorelements, such as Fe, Al, and Ca. For Cd, Zn, Ni, As, Cr, andMn there was a similarity between ammonium-EDTA andacetic acid as an extracting agent. Nevertheless, the yields forCr were negligible.

From Figure 1, it can be deduced that comparableamounts of Zn and Ni are extracted with both reagents.Slightly more Cd was extracted with acetic acid comparedto ammonium-EDTA, while for the trace elements Cuand Pb, ammonium-EDTA was a more effective extracting


Table 4: Average, standard deviation, median, minimum, and maximum of the amounts of major and trace metals, the organic carbon(OC) content, and the pH in the 72 sediments investigated in this study.

Cr Cu Zn Ni As Cd Pb Co Mn K Fe Ca Al OCpH

mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg % % % %

Average 131 65 706 29 80 18 124 7.9 273 12265 6 2 3 6.2 6.7

Stdv 99 71 959 23 92 37 162 5.2 342 7106 5 1 1 4.1 0.8

Median 95 45 442 27 42 9 69 8.0 172 11070 3 2 3 5.3 6.7

Min 28 1 14 3 4 0.1 4 1.3 23 3248 1 0.2 1 0.5 5.3

Max 529 286 5086 117 516 198 756 24.0 2555 46479 23 5 5 19.5 8.1

Table 5: Average, standard deviation, median, minimum, and maximum of the amounts of major and trace metals extracted with aceticacid and ammonium-EDTA.

Acetic acid

Cr Cu Zn Ni As Cd Pb Co Mn K Fe Ca AlpH

mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg

Average 2 10.4 436 8.02 3.26 8 5 1.3 104 338 1049 11386 172 3.1

Stdv 2 15.3 652 8.81 5.60 13 11 1.2 121 786 2223 8865 126 0.4

Median 1 3.3 197 6.34 1.49 5 1 1.1 74 195 296 7846 145 3.1

Min <0.01 <0.01 4 <0.01 0.05 0.1 <0.01 0.1 2 65 47 1126 28 2.2

Max 10 77.0 3574 43.88 26.96 69 65 4.8 592 6788 13143 38694 844 3.6

Ammonium-EDTA

Cr Cu Zn Ni As Cd Pb Co Mn K Fe Ca AlpH

mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg

Average 1 31 419 7.76 2.4 10 49 0.7 50 228 3176 9066 68 6.3

Stdv 1 43 578 9.56 3.6 21 81 0.6 66 143 2957 4478 59 1.0

Median 1 20 216 5.65 1.2 5 21 0.5 41 182 2039 9178 50 6.4

Min <0.01 0.10 2 0.45 0.02 <0.01 <0.01 <0.01 <0.01 51 176 1232 8 4.6

Max 6 177 2539 51.67 16 109 389 3.3 527 603 11902 18718 265 7.8

agent (Figure 1). Some samples with a high total con-tent of Zn and Cd, as well as an important EDTA andacetic acid extractable content of Zn and Cd, distort thegeneral relationship between total and ammonium-EDTAor acetic acid extractable content of Zn and Cd. Log-transformed concentrations were used in further calculations(see Section 3.2).

The more important release of Cu and Pb byammonium-EDTA compared to acetic acid is related tothe fact that EDTA has a much stronger complexing capacitythan acetic acid. The stability constants for EDTA andacetic acid complexes with Cd, Cu, Zn, Ni, and Pb arepresented in Table 7. Complexes of Cd and Zn with acetateare characterised by significantly lower stability constantsthan Pb and Cu, but this does not compensate for thefact that Zn and Cd are more easily mobilised upon a pHdecrease compared to Cu and Pb. The pH of the acetic acidextract is in the range 2.7–3.5, promoting the release of Cd,Zn, and Ni while Cu and Pb are only released in significantamounts at lower pH values.

Since the pH of the ammonium-EDTA extract is fairlyconstant and almost neutral, trace element release is prin-cipally a consequence of complexation reactions. No rela-tionship was found between the formation constants for 1 : 1

EDTA complexes and the slopes of the relation in Figure 1,indicating that the amount of a metal extracted with EDTAis not only determined by the affinity of the metal to formcomplexes with EDTA. Besides the complexing affinity of aEDTA for a trace element, the affinity between the sedimentand a trace element will also control the desorption by thecomplexing agent.

According to Papassiopi et al. [39], the dissolution ofcalcite can consume EDTA in calcareous soils, lowering theextraction efficiency for trace elements. In general, majorcations present in the soil may be one of the factors affectingtrace element extraction efficiency [66]. In the presentstudy, important amounts of Fe and Ca were also extractedwith ammonium-EDTA (Table 5), possibly affecting theextraction efficiency of the reagent.

The comparison of “potential availability” of traceelements by different and single extractions (acetic acidor ammonium-EDTA extractions as defined by the SMTprogram) is not completely straightforward since operationalconditions (L/S ratio, extraction time, etc.) and reagentsare different. The estimation of the potential (long-term)trace element availability is thus operationally defined by theextracting agents used. van der Sloot et al. [51] mention thatthe acetic acid and ammonium-EDTA extractions release


Table 6: Results of the one-way ANOVA for the comparison of elements extracted with acetic acid and EDTA (∗significant at P < 0.05).

SS df MS F P-value

Cr 8.5 1 8.5 3.15 0.078

Cu 15448 1 15448 15.05 0.000∗

Zn 9944 1 9944 0.03 0.872

Ni 2.5 1 2.5 0.03 0.863

As 26.1 1 26.1 1.18 0.279

Cd 137.2 1 137.2 0.45 0.502

Pb 71795 1 71795 21.38 0.000∗

Co 15.2 1 15.2 18.28 0.000∗

Mn 104719 1 104719 11.05 0.001∗

Fe 162860447 1 162860447 23.80 0.000∗

Ca 193751031 1 193751031 3.93 0.049∗

Al 387317 1 387317 39.76 0.000∗

Table 7: Formation constants (log K) for 1 : 1 metal EDTA complexes and 1 : 1 metal acetate complexes.

Stability constant for acetic acid Reference Stability constant for EDTA Reference

Cu2+ 1.76, 1.87, 2.71, 3.09, 3.63 [61, 62] 19.7, 18.8 [63, 64]

Zn2+ 0.91, 1.9 [61, 62] 17.5, 16.5 [63, 64]

Ni2+ 0.74, 0.83, 1.43 [61, 62] 19.5, 18.56 [63, 64]

Cd2+ 1.19, 1.23, 1.32, 1.82, 3.15 [61, 62] 17.4 [63, 64]

Pb2+ 3.5, 2.98, 4.08 [61, 62] 19 [63, 64]

Co2+ 0.71 [65] 16.21 [64]

Mn2+ 13.56 [64]

Fe3+ 25.7 [64]

Ca2+ 0.5 [65] 10.7 [64]

Al3+ 16.3 [64]

an amount of Zn and Cd that is very similar to the amountleached at pH 4 in a 24 h pHstat test, L/S = 5 L/kg). However,this seems to not be the case in a more heterogeneousdataset with samples with very different physicochemicalcharacteristics.

The difference between acetic acid and ammonium-EDTA as an extracting agent is particularly interesting for themetalloid As. Although As is generally more easily mobilizedwhen pH increases, acetic acid is capable of extracting aconsiderable amount of As. Even more As is extracted byacetic acid than with ammonium-EDTA, despite the lowerfinal pH of the acetic acid extract (around Section 3.5).Wenzel et al. [67] also found that only minor proportionsof As were extracted by ammonium-EDTA, virtually notcontributing to As fractionation. The inefficiency of Asextraction by ammonium-EDTA is explained by the fact thatEDTA does not form stable complexes with arsenic [68].

3.4. Influence of Sediment Characteristics on Metal Extractabil-ity. The lack of a distinct relationship between total andacid extractable metal concentrations in the present datasetis most likely due to the very different physicochemical andmineralogical characteristics of the samples. In a previousstudy on land-disposed dredged sediments [69], acetic acid

(0.43 mol/L) extractable Zn and Cd concentrations werelinearly correlated with total Zn and Cd concentrations. Allthe samples originated from the same river catchment andwere characterised by and elevated clay and organic carboncontent.

The pH after extractions with acetic acid (0.43 mol/L)was in the range 2.2–3.6. The increase in pH after extraction(Figure 2) can be related to the acid neutralizing capacity ofthe samples. The pH of the EDTA extract was between 4.6and 7.8, despite the fact that the ammonium-EDTA solutionis a buffered solution. Additionally, the difference in pH ofthe acetic acid extracts before and after extraction showeda significant linear correlation with the (log-transformed)total Ca content of the samples. This can simply be explainedby the fact that samples with an elevated acid neutralizingcapacity (ANC) usually are characterised by a high CaCO3

content, which is an important contributor to acid neutraliz-ing reactions in soils and sediments.

Based on the results of the extractions and on the sed-iment composition, regression equations were constructedusing the ammonium-EDTA or acetic acid extractablecontent of a trace element (Cu, Cr, Zn, Ni, As, Cd, Pb, orCo) as the dependent variable and the total content of therespective trace element (Cu, Cr, Zn, Ni, As, Cd, Pb, or


Ext

ract

ed c

once

ntr

atio

n (

mg/

kg) 12

10

8

6

4

2

00 200 400 600

y = 0.002x + 1.0792

R2 = 0.0418

y = 0.0147x − 0.1034

R2 = 0.4686

Cr

Total concentration (mg/kg)

(a)

Ext

ract

ed c

once

ntr

atio

n (

mg/

kg)


200

180

160

140

120

100

80

60

40

20

00 100 200 300 400

y = 0.5319x − 3.3676

R2 = 0.7823

y = 0.1776x − 1.1343R2 = 0.6789

Cu

(b)

4000

3500

3000

2500

2000

1500

1000

500

00 2000 4000 6000

y = 0.5363x + 56.957R2 = 0.6226

y = 0.4828x + 78.119R2 = 0.6413

Zn

Ext

ract

ed c

once

ntr

atio

n (

mg/

kg)


(c)

Ext

ract

ed c

once

ntr

atio

n (

mg/

kg) 60

50

40

30

20

10

00 50 100 150

y = 0.3241x − 1.3724

R2 = 0.6907

y = 0.3399x − 2.0968R2 = 0.6454

Ni


(d)

30

25

20

15

10

5

0

y = 0.024x + 1.3485R2 = 0.1557

y = 0.0163x + 1.1062R2 = 0.1753

As

Ext

ract

ed c

once

ntr

atio

n (

mg/

kg)

0 200 400 600


(e)

120

100

80

60

40

20

0

y = 0.5489x − 0.2674R2 = 0.9681

y = 0.3434x + 1.4465R2 = 0.9348

Cd

Ext

ract

ed c

once

ntr

atio

n (

mg/

kg)


0 100 200 300

(f)

450

400

350

300

250

200

150

100

50

00 200 400 600 800

y = 0.4843x − 10.854R2 = 0.9352

y = 0.0567x − 2.5234R2 = 0.7055

Pb

Ext

ract

ed c

once

ntr

atio

n (

mg/

kg)


(g)

5

Co6

4

2

0

1

3

Ext

ract

ed c

once

ntr

atio

n (

mg/

kg)

3020100

y = 0.1645x + 0.0268R2 = 0.5497

y = 0.0181x + 0.5381R2 = 0.0257


(h)

Figure 1: Amount of trace elements extracted by ammonium-EDTA (black symbols and full line) and acetic acid (white symbols and dottedline) as a function of their total content in the sediments.


−0.5 0 0.5 1 1.5

pH difference

2

2.5

3

3.5

4

4.5

5

Log

(Ca)

y = 0.8348x + 3.3264

R2 = 0.7081

Figure 2: Difference between the pH of the acetic acid solutionbefore and after extraction versus the amount of Ca released fromthe samples.

Co), the total content of Ca, Al, and Fe, the pH, and theorganic carbon content as independent variables. Becausethe data were not normally distributed, log-transformed datawere used (expect for pH, which is already a logarithmicvalue). In soils and sediments, Ca is dominantly found inclay minerals and in carbonate minerals (e.g., CaCO3). Fe is amajor component of Fe(hydr)oxides, but is also a constituentof clay minerals. Al is often used as a proxy for the claycontent in soils and sediments. Rodrigues et al. [70] foundthat the variation of the total Al in soil content in soil andsediments expresses the sorptive capacity of aluminosilicatesand Al oxides at the surfaces and edges of clay minerals betterthan the actual variability of clay contents.

For most trace elements, the independent variables (totalconcentrations of trace metals, Al, Fe, Ca, organic carbon(OC), and/or (pH)) were significant in predicting the depen-dent variable (ammonium-EDTA or acetic acid extractableconcentration) when the level of significance is below 0.05.For almost all trace elements, the total content was the mostimportant predicting variable. For the ammonium-EDTAextractable content of Zn, Ni, Cd, and Pb, the organic carboncontent was also an important predicting variable, which canbe related to the fact that EDTA is able to form organometalcomplexes, which compete with organic matter in soil. Thepredictive value of total Al and Fe concentrations differedamong different trace elements.

For Zn, Ni, As, and Cd, the total Ca content was anexplaining variable for their acid extractable content, besidesthe total content of these elements, whereas the predictivevalue of Fe and Al was less important.

In the present study, the “EDTA extractable pool” ofCu was not significantly affected by the major elementsand organic carbon content of the sediments since theEDTA extractable fraction of this element was found to beproportional to its total content (Table 8). For the amountof Cu extracted with acetic acid, pH was also a significantexplaining variable, together with the total Cu content of thesediments.

A linear correlation between EDTA extractable and totalconcentrations of Cd, Zn, Cu, and Pb has been mentionedin several other studies [71, 72]. McGrath [73] mentionedthat the amount of Cd, Cu, and Ni extracted by EDTA from7 polluted and unpolluted Irish soils was related to totalmetal content, while the EDTA extractable Zn content wasrelated to organic carbon. Filipek and Pawlowski [74] alsomentioned a positive relationship between EDTA extractableCu and organic carbon in soils, a relationship not observedin our dataset.

3.5. Potential Use and Interpretation of

the Results of Ammonium-EDTA and Acid

Extractable Metal Concentrations

3.5.1. Quantification of Metal Partitioning in Soils. Metalpartitioning in soils can be quantified by models in whichmetal concentrations in the pore water are described as afunction of the metal binding solid phases such as Fe and Al(hydr)oxides, organic matter, and clay and as a function ofsoil characteristics that influence trace element partitioning,such as pH. The ratio between total metal content boundto a soil relative to its concentration in the soil solutionis often represented by Kd coefficients. However, such amodel assumes that the sorption capacity of a material isindependent of the soil properties (organic matter content,pH, clay content, etc.) and, therefore, single Kd valuesare not appropriate to predict metal solubility in soil. Inthe literature, several models can be found, in which it isassumed that exchangeable metals and protons compete foradsorption on soil exchange sites [75]. In these models,besides soil properties—such as pH, organic carbon content,and CEC—total elements concentrations are often usedto predict dissolved elements concentrations (i.e., elementsconcentrations in the pore water). However, several authors[76, 77] used extractions with HNO3 (0.43 mol/L), HCl, orEDTA to estimate the available pool of an element for use insolid-solution partitioning modeling, instead of using totalmetal concentrations.

In a study of Cappuyns et al. [78], ammonium-EDTAextractable metal concentrations were used instead of totalmetal concentrations. Stepwise multiple linear regressionwas performed with pH(CaCl2), organic carbon content, claycontent, CaCl2 extractable element concentrations, dissolvedorganic carbon (DOC) content in de CaCl2 extracts, andammonium-EDTA extractable element concentrations. Withthis model, in which the EDTA extractable element fractionrepresented the “available” pool of the element, Cd and Znconcentrations in the pore water could be predicted very well,based on pH, organic carbon content, and EDTA extractableZn and Cd concentrations. Rodrigues et al. [79] derivedFreundlich-type models based on commonly available soilproperties (pH, organic carbon, and clay) as well as extendedmodels that used other properties such as amorphous Al andFe oxides and evaluated their possible use in risk assessment.The approach enabled the prediction of the reactivity ofpotentially toxic elements.

Empirical partition relations can be useful to de-scribe metal partitioning in soils, especially in large-scale


Table 8: Ammonium-EDTA and acetic acid extractable element concentrations as a function of their total concentrations and totalconcentrations of Al, Fe, Ca, and organic carbon (OC). (n = 72, ∗significant at α = 0.05). Acetic acid is abbreviated as “HOAc” in theequations.

Acetic acid r2

log [Cu]HOAc = 1.073 log [Cu]total + 0.363 pH − 3.645 0.851∗

log [Cr]HOAc = 0.958 log [Cr]total− 0.408 log [Fe]total− 1.683 0.345

log [Zn]HOAc = 1.013 log [Zn]total + 0.479 log [Ca]total− 0.399 log [Al]total− 0.312 0.856∗

log [Ni]HOAc = 0.915 log [Ni]total + 0.557 log [Ca]total− 0.439 log [Al]total− 0.478 0.766∗

log [As]HOAc = 1.356 log [Ca]total + 0.649 log [As]total− 0.617 log [Fe]total + 0.641 log [Al]total− 1.101 0.751∗

log [Cd]HOAc = 0.663 log [Cd]total + 0.268 log [Ca]total− 0.077 0.835∗

log [Pb]HOAc = −2.93 pH + 0.75 log [Pb]total− 0.905 log [Fe]total + 1.071 log [Al]total− 0.598 log [OC] − 1.289 0.968∗

log [Co]HOAc = 0.591 log [Co]total + 0.457 log [Ca] − 0.591 0.590∗

Ammonium-EDTA

log [Cu]EDTA = 1.058 log [Cu]total− 0.546 0.698∗

log [Cr]EDTA = 0.512 log [Cr]total + 0.159 pH − 0.063 0.393

log [Zn]EDTA = 1.043 log [Zn]total + 0.368 log [OC] + 0.268 log [Ca]total− 0.765 0.883∗

log [Ni]EDTA = 0.858 log [Ni]total± 0.172 pH + 0.326 log [OC] + 0.473 0.715∗

log [As]EDTA = 0.878 log [As]total + 0.513 log [Al]total− 0.748 log [Fe]total + 0.373 log [Ca]total− 1.616 0.670∗

log [Cd]EDTA = 0.663 log [Cd]total + 0.345 log [OC] + 0.220 log [Fe]total− 0.62 0.883∗

log [Pb]EDTA = 1.017 log [Pb]total− 0.338 log [Al]total + 0.166 log [Fe]total + 0.183 log [OC] + 0.085 pH − 1.289 0.749∗

log [Co]EDTA = −0.384 log [OC] + 0. 333 log [Co]total + 3.75 0.192

applications, and offer the advantage that they are simple tounderstand and can be used by nonexpert users [80].

3.5.2. Relation between EDTA Extractable Metal Concentra-tions and Plant Uptake? For trace elements, total concentra-tions in soil are mostly used as an input in risk assessmentmodels. However, trace elements in contaminated soil arerarely released completely as only a portion of trace elementsis “bioavailable” or “geoavailable” (which means that theelements can be released and become available for biologicaluptake). Several authors claim that extractions such asextractions with EDTA and acetic acid can be used toevaluate the bioavailable fraction of an element in soil.However, the interpretation of the term “bioavailability” canvary widely as the availability and uptake of elements will alsodepend of the type of organism (plant, earthworm). EDTAsalts have mainly been used to assess the availability of metalsto plants, since they are believed to mimic rhizosphere effectsin the soil. For example, a study of Bakircioglu et al. [81]indicated the extractable Pb and Ni of soils by EDTA singleextraction procedures were significantly correlated with themetal contents of wheat grains. According to Anyanwuet al. [82], EDTA can be used to quantify the empiricalrelationships between plant uptake and soil metal contents.In another study, a significant positive correlation was foundbetween EDTA extractable metals and metal accumulation inthe shoot of Brassica juncea L. [83]. However, other studies(e.g., [84]) showed that extracting agents such as EDTA cannot be used as universal soil extraction for estimating Cu, Zn,and Ni uptake by barley.

Peijnenburg et al. [85] made an overview of empiricalmethods for extraction of metals from soils as a surrogate forbioavailable and bioaccessible metal pools. They concluded

that the value of these chemical methods for measuringbioavailability can be significantly improved when thespecies, metal, and soil specific aspects of bioavailability aremore accurately taken into account.

3.5.3. Estimation of “Available Element Fraction” for Use inRisk Assessment and LCA. In the case of remediation projectsfor soils and sediments, trace element “availability” is alsoan important consideration since only a portion of thetotal metal load in soils or sediments can be considered as“geoavailable” or mobile. In a risk-conservative approach,it is assumed that all the metals contained in a solidmatrix (soil, sediment, waste material, etc.) will be released.When this “total metal content” is used as an input inlife cycle analysis (LCA), this will most likely result in anoverestimation of the risk associated with the trace elements.Additionally, LCA typically covers an extended period oftime (depending on the life cycle of a product or process),so long-term trace element emissions have to be assessed.The assessment of trace element released from soils andsediments on the long term is still controversial. Severalmethods and procedures have been proposed to estimatethe long-term emissions of trace elements contained in soils,sediments, and waste materials, but there is no consensus asto which method performs best [86]. Extractions with “mild”reagents such as ammonium-EDTA and acetic acid can beuseful to estimate “geoavailable” element concentrations insoils and sediments, and they have the advantage that theyare relatively easy to perform.

4. Conclusion

The present study shows that both ammonium-EDTA andacetic acid have a different capacity to extract elements,


depending on the elements and on the characteristics ofthe matrix that is being extracted (sediments in this case).The estimation of the potential (long-term) trace elementavailability is operationally defined by the extracting agentsused. Making the comparison of “potential availability”by different single extractions (acetic acid or ammonium-EDTA extractions as defined by the SMT program) isnot completely straightforward since operational condi-tions (L-S ratio, extraction time, etc.) and reagents aredifferent.

For most trace elements (Cu, Zn, Cd, Ni, As, andPb), a significant linear relationship could be establishedbetween their ammonium-EDTA or acetic acid extractableconcentrations and their total concentrations, organic car-bon content pH, and Fe, Al, and/or Ca content. The scientificunderstanding of trace element partitioning in the complexsoil-water system with these simple models is rather limited,but they offer the opportunity to use data from singleextractions in a more comprehensive way.

Although single extractions with extracting agents—such as ammonium-EDTA and acetic acid that aim toextract the “mobilizable” pool of elements from soils andsediments—can be statistically related to soil properties,they are primarily determined by the total element contentin the soil or sediments. Despite the fact that singleextractions with ammonium-EDTA and acetic acid cannotdirectly be related to the bioavailability of elements, theycan provide input data for use in risk assessment models.Additionally, they also offer possibilities to perform a fastscreening of the mobilizable pool of elements in soils andsediments.

Acknowledgments

Grateful acknowledgements are made to Talia Stough for theimprovement of the English language of the paper and tothe anonymous reviewers for their constructive commentson the first version of the paper.

References

[1] J. H. Duffus, “Heavy metals—a meaningless term?” Pure andApplied Chemistry, vol. 74, no. 5, pp. 793–807, 2002.

[2] M. E. Hodson, “Heavy metals—geochemical bogey men?”Environmental Pollution, vol. 129, no. 3, pp. 341–343, 2004.

[3] P. M. Driessen and R. Dudal, The Major Soils of the World, Lec-ture Notes on Their Geography, Formation, Proiperties andUse, Agricultural University of Wageningen and KatholiekeUniversiteit Leuven, 1991.

[4] OVAM, Openbare Vlaamse Afvalstoffenmaatschappij. Ontwerpuitvoeringsplan bagger- en ruimingspecie, OVAM, Mechelen,Belgium, 2007.

[5] IUSS Working Group WRB, “World reference base for soilresources A framework for international classification, cor-relation and communication,” World Soil Resources Reports103, FAO, Rome, Italy, 2006.

[6] A. C. Scheinost, R. Kretzschmar, S. Pfister, and D. R.Roberts, “Combining selective sequential extractions, X-rayabsorption spectroscopy, and principal component analysis

for quantitative zinc speciation in soil,” Environmental Scienceand Technology, vol. 36, no. 23, pp. 5021–5028, 2002.

[7] W. J. G. M. Peijnenburg, M. Zablotskaja, and M. G. Vijver,“Monitoring metals in terrestrial environments within abioavailability framework and a focus on soil extraction,”Ecotoxicology and Environmental Safety, vol. 67, no. 2, pp. 163–179, 2007.

[8] A. L. Nolan, M. J. McLaughlin, and S. D. Mason, “Chemicalspeciation of Zn, Cd, Cu, and Pb in pore waters of agriculturaland contaminated soils using donnan dialysis,” EnvironmentalScience and Technology, vol. 37, no. 1, pp. 90–98, 2003.

[9] J. L. Roulier, S. Belaud, and M. Coquery, “Comparison ofdynamic mobilization of Co, Cd and Pb in sediments usingDGT and metal mobility assessed by sequential extraction,”Chemosphere, vol. 79, no. 8, pp. 839–843, 2010.

[10] M. Koster, L. Reijnders, N. R. Van Oost, and W. J. G.M. Peijnenburg, “Comparison of the method of diffusivegels in thin films with conventional extraction techniquesfor evaluating zinc accumulation in plants and isopods,”Environmental Pollution, vol. 133, no. 1, pp. 103–116, 2005.

[11] H. Vandenhove, K. Antunes, J. Wannijn, L. Duquene, and M.Van Hees, “Method of diffusive gradients in thin films (DGT)compared with other soil testing methods to predict uraniumphytoavailability,” Science of the Total Environment, vol. 373,no. 2-3, pp. 542–555, 2007.

[12] O. Mojsilovic, R. G. McLaren, and L. M. Condron, “Modellingarsenic toxicity in wheat: simultaneous application of diffusivegradients in thin films to arsenic and phosphorus in soil,”Environmental Pollution, vol. 159, no. 10, pp. 2996–3002, 2011.

[13] P. M. V. Nirel and F. M. M. Morel, “Pitfalls of sequentialextractions,” Water Research, vol. 24, no. 8, pp. 1055–1056,1990.

[14] I. Maiz, M. V. Esnaola, and E. Millan, “Evaluation of heavymetal availability in contaminated soils by a short sequentialextraction procedure,” Science of the Total Environment, vol.206, no. 2-3, pp. 107–115, 1997.

[15] J. L. Gomez Ariza, I. Giraldez, D. Sanchez-Rodas, and E.Morales, “Comparison of the feasibility of three extractionprocedures for trace metal partitioning in sediments fromsouth-west Spain,” Science of the Total Environment, vol. 246,no. 2-3, pp. 271–283, 2000.

[16] P. S. Rendell, G. E. Batley, and A. J. Cameron, “Adsorptionas a control of metal concentrations in sediment extracts,”Environmental Science and Technology, vol. 14, no. 3, pp. 314–318, 1980.

[17] J. L. Howard and W. J. Vandenbrink, “Sequential extractionanalysis of heavy metals in sediments of variable compositionusing nitrilotriacetic acid to counteract resorption,” Environ-mental Pollution, vol. 106, no. 3, pp. 285–292, 1999.

[18] A. Tessier and P. G. C. Campbell, “Comments on the testingof the accuracy of an extraction procedure for determiningthe partitioning of trace metals in sediments,” AnalyticalChemistry, vol. 60, no. 14, pp. 1475–1476, 1988.

[19] A. Tessier and P. G. C. Campbell, “Comment on ’Pitfalls ofsequential extractions’ by P.M.V. Nirel and F.M.M. Morel,”Water Research, vol. 24, pp. 1055–1056, 1991.

[20] E. Tipping, N. B. Hetherington, J. Hilton, D. W. Thompson,E. Bowles, and J. Hamilton-Taylor, “Arrtifacts in the useof chemical extraction to determine distributions of metalsbetween oxides of manganese and iron,” Analytical Chemistry,vol. 57, pp. 1944–1946, 1985.

[21] K. A. Gruebel, J. A. Davis, and J. O. Leckie, “The feasibilityof using sequential extraction techniques for arsenic and


selenium in soils and sediments,” Soil Science Society ofAmerica Journal, vol. 52, no. 2, pp. 390–397, 1988.

[22] R. T. Dhoum and G. J. Evans, “Evaluation of uraniumand arsenic retention by soil from a low level radioactivewaste management site using sequential extraction,” AppliedGeochemistry, vol. 13, no. 4, pp. 415–420, 1998.

[23] M. J. La Force and S. Fendorf, “Solid-phase iron characteri-zation during common selective sequential extractions,” SoilScience Society of America Journal, vol. 64, no. 5, pp. 1608–1615, 2000.

[24] X.-Q. Shan and B. Chen, “Evaluation of sequential extractionfor speciation of trace metals in model soil containing naturalminerals and humic acid,” Analytical Chemistry, vol. 65, no. 6,pp. 802–807, 1993.

[25] J. L. Gomez Ariza, I. Giraldez, D. Sanchez-Rodas, and E.Morales, “Selectivity assessment of a sequential extractionprocedure for metal mobility characterization using modelphases,” Talanta, vol. 52, no. 3, pp. 545–554, 2000.

[26] S. Van Herreweghe, R. Swennen, C. Vandecasteele, and V.Cappuyns, “Solid phase speciation of arsenic by sequentialextraction in standard reference materials and industriallycontaminated soil samples,” Environmental Pollution, vol. 122,no. 3, pp. 323–342, 2003.

[27] P. Quevauviller, “Operationally defined extraction proceduresfor soil and sediment analysis I. Standardization,” Trends inAnalytical Chemistry, vol. 17, no. 5, pp. 289–298, 1998.

[28] A. Ure, P. Quevauviller, H. Munteau, and B. Griepink,“Improvements in the determination of extractable contentsof trace metals in soils and sediments prior to certification,”Tech. Rep., Community Bureau of reference, Commission ofthe European Communities, 1993.

[29] G. Rauret, J. F. Lopez-Sanchez, A. Sahuquillo et al., “Improve-ment of the BCR three step sequential extraction procedureprior to the certification of new sediment and soil referencematerials,” Journal of Environmental Monitoring, vol. 1, no. 1,pp. 57–61, 1999.

[30] A. Sahuquillo, J. F. Lopez-Sanchez, R. Rubio et al., “Use ofa certified reference material for extractable trace metals toassess sources of uncertainty in the BCR three-stage sequentialextraction procedure,” Analytica Chimica Acta, vol. 382, no. 3,pp. 317–327, 1999.

[31] S. K. Gupta, M. K. Vollmer, and R. Krebs, “The importanceof mobile, mobilisable and pseudo total heavy metal fractionsin soil for three-level risk assessment and risk management,”Science of the Total Environment, vol. 178, pp. 11–20, 1996.

[32] W. P. Miller, D. C. Martens, L. W. Zelazny, and E. T. Kornegay,“Forms of solid phase copper in copper-enriched swinemanure,” Journal of Environmental Quality, vol. 15, no. 1, pp.69–72, 1986.

[33] O. K. Borggaard, “Selective extraction of amorphous ironoxides by EDTA from a Danish sandy loam,” Journal of SoilScience, vol. 30, pp. 727–734, 1979.

[34] B. Nowack and L. Sigg, “Dissolution of Fe(III)(hydr) oxidesby metal-EDTA complexes,” Geochimica et Cosmochimica Acta,vol. 61, no. 5, pp. 951–963, 1997.

[35] M. A. M. Kedziorek and A. C. M. Bourg, “Solubilization oflead and cadmium during the percolation of EDTA througha soil polluted by smelting activities,” Journal of ContaminantHydrology, vol. 40, no. 4, pp. 381–392, 2000.

[36] L. Madrid and E. Dıaz-Barrientos, “Release of metals fromhomogeneous soil columns by wastewater from an agricul-tural industry,” Environmental Pollution, vol. 101, no. 1, pp.43–48, 1998.

[37] B. Sun, F. J. Zhao, E. Lombi, and S. P. McGrath, “Leachingof heavy metals from contaminated soils using EDTA,”Environmental Pollution, vol. 113, no. 2, pp. 111–120, 2001.

[38] Z. Zou, R. Qiu, W. Zhang et al., “The study of operatingvariables in soil washing with EDTA,” Environmental Pollution,vol. 157, no. 1, pp. 229–236, 2009.

[39] N. Papassiopi, S. Tambouris, and A. Kontopoulos, “Removalof heavy metals from calcareous contaminated soils by EDTAleaching,” Water, Air, and Soil Pollution, vol. 109, no. 1–4, pp.1–15, 1999.

[40] N. Finzgar and D. Lestan, “Multi-step leaching of Pb and Zncontaminated soils with EDTA,” Chemosphere, vol. 66, no. 5,pp. 824–832, 2007.

[41] R. S. Tejowulan and W. H. Hendershot, “Removal of tracemetals from contaminated soils using EDTA incorporatingresin trapping techniques,” Environmental Pollution, vol. 103,no. 1, pp. 135–142, 1998.

[42] P. Theodoratos, N. Papassiopi, T. Georgoudis, and A. Kon-topoulos, “Selective removal of lead from calcareous pollutedsoils using the Ca-EDTA salt,” Water, Air, and Soil Pollution,vol. 122, no. 3-4, pp. 351–368, 2000.

[43] A. Bermond and J. P. Ghestem, “Kinetic study of trace metalEDTA-desorption from contaminated soils,” in Heavy MetalsRelease in Soils, H. M. Selim and D. L. Sparks, Eds., pp. 131–147, Lewis, 2001.

[44] A. Bermond, “Limits of sequential extraction procedures re-examined with emphasis on the role of H+ ion reactivity,”Analytica Chimica Acta, vol. 445, no. 1, pp. 79–88, 2001.

[45] A. Polettini, R. Pomi, E. Rolle et al., “A kinetic study of chelant-assisted remediation of contaminated dredged sediment,”Journal of Hazardous Materials, vol. 137, no. 3, pp. 1458–1465,2006.

[46] D. F. Kalf, M. A. G. T. Hoop van den, J. P. Rila, C.Posthuma, and T. P. Traas, “Environmental Risk Limits forEthylene Diamine Tetra Acetic acid (EDTA). Rijksinstituutvoor Volksgezondheid en Milieu RIVM (Ed.),” RIVM Report601501010/2003, Bilthoven, The Netherlands, 2003.

[47] D. Jiraroj, F. Unob, and A. Hagege, “Degradation of Pb-EDTAcomplex by a H2O2/UV process,” Water Research, vol. 40, no.1, pp. 107–112, 2006.

[48] M. Pociecha and D. Lestan, “EDTA leaching of Cu contam-inated soil using electrochemical treatment of the washingsolution,” Journal of Hazardous Materials, vol. 165, no. 1–3,pp. 533–539, 2009.

[49] A. Tessier, P. G. C. Campbell, and M. Blsson, “Sequentialextraction procedure for the speciation of particulate tracemetals,” Analytical Chemistry, vol. 51, no. 7, pp. 844–851, 1979.

[50] A. M. Ure, “Single extraction schemes for soil analysis andrelated applications,” Science of the Total Environment, vol. 178,pp. 3–10, 1996.

[51] H. A. van der Sloot, R. N. J. Comans, and O. Hjelmar,“Similarities in the leaching behaviour of trace contaminantsfrom waste, stabilized waste, construction materials and soils,”Science of the Total Environment, vol. 178, pp. 111–126, 1996.

[52] D. W. Nelson and L. E. Sommers, “Total carbon, organiccarbon and organic matter,” in Methods of Soil Analysis, part 2:Chemical and Biological Properties, pp. 516–593, 2nd edition,198.

[53] R. Chhabra, J. Pleysier, and A. Cremers, “The measurementof the cation exchange capacity and exchangeable cations insoils: a new method,” in Proceedings of the International ClayConference, S. W. Bailey, Ed., pp. 439–449, Applied PublishingLtd, Wilmette, Ill, USA, 1975.


[54] L. P. van Reeuwijk, Procedures for Soil Analysis, ISRIC,Wageningen, The Netherlands, 3rd edition, 1992.

[55] C. A. Brockhoff, J. T. Creed, T. D. Martin, E. R. Martin, andS. E. Long, “EPA Method 200.8, Revision 5.5: determinationof trace metals in waters and wastes by inductively coupledplasma-mass spectrometry,” EPA-821R-99-017, p. 61 , Octo-ber 1999.

[56] R. Webster, “Regression and functional relations,” EuropeanJournal of Soil Science, vol. 48, no. 3, pp. 557–566, 1997.

[57] R. Webster, “Statistics to support soil research and theirpresentation,” European Journal of Soil Science, vol. 52, no. 2,pp. 331–340, 2001.

[58] P. Quevauviller, E. A. Maier, B. Griepink, U. Fortunati, K.Vercoutere, and H. Muntau, “Certified reference materialsof soils and sewage sludges for the quality control of tradeelement environmental monitoring,” Trends in AnalyticalChemistry, vol. 15, no. 10, pp. 504–513, 1996.

[59] M. Pueyo, G. Rauret, D. Luck et al., “Certification of theextractable contents of Cd, Cr, Cu, Ni, Pb and Zn in afreshwater sediment following a collaboratively tested andoptimised three-step sequential extraction procedure,” Journalof Environmental Monitoring, vol. 3, no. 2, pp. 243–250, 2001.

[60] P. Quevauviller, G. Rauret, A. Ure, J. Bacon, and H. Muntau,“The certification of the EDTA and acetic acid-extractablecontents (mass fractions) of Cd, Cr, Cu, Pb and Zn insewage sludge amended soils CRMs 484 and 484,” EuropeanCommision, BCR information, Reference materials. EUR17127 EN, p. 99, 1997.

[61] A. E. Martell and R. M. Smith, Critical Stability Constants.Other Organic Ligands, vol. 3, Plenum Press, New York, NY,USA, 1977.

[62] A. E. Martell and R. M. Smith, Critical Stability Constants. FirstSupplement, vol. 5, Plenum Press, New York, NY, USA, 1982.

[63] W. A. Norvell, “Reactions of metal chelates in soil and nutrientsolutions,” in Microelements in Agriculture, J. J. Mortvedt, F. R.Cox, L. M. Shuman, and R. M. Welch, Eds., pp. 187–227, SoilScience Society of America, Madison, Wis, USA, 2nd edition,1991.

[64] T. E. Furia, “Chapter 6—sequestrants in foods,” in CRCHandbook of Food Additives, pp. 271–294, CRC Press, WestPalm Beach, Fla, USA, 2nd edition, 1972.

[65] J. W. Buthing and K. M. Thong, “Stability constants forsome 1 : 1 metal-carboxylate complexes,” Canadian Journal ofChemistry, vol. 48, pp. 1654–1656, 1970.

[66] C. Kim, Y. Lee, and S. K. Ong, “Factors affecting EDTA extrac-tion of lead from lead-contaminated soils,” Chemosphere, vol.51, no. 9, pp. 845–853, 2003.

[67] W. W. Wenzel, N. Kirchbaumer, T. Prohaska, G. Stingeder,E. Lombi, and D. C. Adriano, “Arsenic fractionation in soilsusing an improved sequential extraction procedure,” AnalyticaChimica Acta, vol. 436, no. 2, pp. 309–323, 2001.

[68] S. Tokunaga and T. Hakuta, “Acid washing and stabilization ofan artificial arsenic-contaminated soil,” Chemosphere, vol. 46,no. 1, pp. 31–38, 2002.

[69] V. Cappuyns, Heavy metal behaviour in overbank sediments andassociated soils, Ph.D. thesis, Katholieke Universiteit Leuven,2004.

[70] S. M. Rodrigues, B. Henriques, E. F. da Silva et al., “Evaluationof an approach for the characterization of reactive andavailable pools of 20 potentially toxic elements in soils: partII—Solid-solution partition relationships and ion activity insoil solutions,” Chemosphere, vol. 81, no. 11, pp. 1560–1570,2010.

[71] S. P. Singh, F. M. G. Tack, and M. G. Verloo, “Solid-phase distribution of heavy metals as affected by singlereagent extraction in dredged sediment derived surface soils,”Chemical Speciation and Bioavailability, vol. 8, no. 1-2, pp. 37–43, 1996.

[72] G. Du Laing, N. Bogaert, F. M. G. Tack, M. G. Verloo,and F. Hendrickx, “Heavy metal contents (Cd, Cu, Zn) inspiders (Pirata piraticus) living in intertidal sediments ofthe river Scheldt estuary (Belgium) as affected by substratecharacteristics,” Science of the Total Environment, vol. 289, no.1–3, pp. 71–81, 2002.

[73] D. McGrath, “Application of single and sequential extractionprocedures to polluted and unpolluted soils,” Science of theTotal Environment, vol. 178, pp. 37–44, 1996.

[74] T. Filipek and L. Pawlowski, “Total and extractable heavy metalcontent of some soils of the Lublin coal mining region,” Scienceof the Total Environment, vol. 96, no. 1-2, pp. 131–137, 1990.

[75] S. Sauve, W. Hendershot, and H. E. Allen, “Solid-solutionpartitioning of metals in contaminated soils: dependence onpH, total metal burden, and organic matter,” EnvironmentalScience and Technology, vol. 34, no. 7, pp. 1125–1131, 2000.

[76] P. F. Romkens, H. Y. Guo, C. L. Chu, T. S. Liu, C. F. Chiang, andG. F. Koopmans, “Characterization of soil heavy metal pools inpaddy fields in Taiwan: chemical extraction and solid-solutionpartitioning,” Journal of Soils and Sediments, vol. 9, no. 3, pp.216–228, 2009.

[77] T. J. Schroder, T. Hiemstra, J. P. M. Vink, and S. E. A. T. M. VanDer Zee, “Modeling of the solid-solution partitioning of heavymetals and arsenic in embanked flood plain soils of the riversrhine and meuse,” Environmental Science and Technology, vol.39, no. 18, pp. 7176–7184, 2005.

[78] V. Cappuyns, R. Swennen, and J. Verhulst, “Assessmentof heavy metal mobility in dredged sediments: porewateranalysis, single and sequential extractions,” Soil and SedimentContamination, vol. 15, no. 2, pp. 169–186, 2006.

[79] S. M. Rodrigues, B. Henriques, E. F. da Silva, M. E. Pereira,A. C. Duarte, and P. F. A. M. Romkens, “Evaluation of anapproach for the characterization of reactive and availablepools of twenty potentially toxic elements in soils: part I—Therole of key soil properties in the variation of contaminants’reactivity,” Chemosphere, vol. 81, no. 11, pp. 1549–1559, 2010.

[80] J. E. Groenenberg, Evaluation of models for metal partitioningand speciation in soils and their use in risk assessment Thesis,Ph.D. thesis, Wageningen University, 2011.

[81] D. Bakircioglu, Y. Bakircioglu-Kurtulus, and H. Ibar, “Com-parison of extraction procedures for assessing soil metalbioavailability of to wheat grains,” Clean, vol. 39, no. 8, pp.728–734, 2011.

[82] E. C. Anyanwu, K. Ijeoma, J. E. Ehiri, and M. A. Saleh,“Bioavailable” Lead Concentration in Vegetable Plants grownin Soil from a reclaimed industrial site: health implications,”Internet Journal of Food Safety, vol. 6, pp. 31–34, 2005.

[83] A. K. Gupta and S. Sinha, “Assessment of single extractionmethods for the prediction of bioavailability of metals toBrassica juncea L. Czern. (var. Vaibhav) grown on tannerywaste contaminated soil,” Journal of Hazardous Materials, vol.149, no. 1, pp. 144–150, 2007.

[84] J. M. Soriano-Disla, I. Gomez, J. Navarro-Pedreno, and A. Lag-Brotons, “Evaluation of single chemical extractants for theprediction of heavy metal uptake by barley in soils amendedwith polluted sewage sludge,” Plant and Soil, vol. 327, no. 1,pp. 303–314, 2010.

[85] W. J. G. M. Peijnenburg, M. Zablotskaja, and M. G.Vijver, “Monitoring metals in terrestrial environments within


a bioavailability framework and a focus on soil extraction,”Ecotoxicology and Environmental Safety, vol. 67, no. 2, pp. 163–179, 2007.

[86] J. Pettersen and E. G. Hertwich, “Critical review: life-cycleinventory procedures for long-term release of metals,” Envi-ronmental Science and Technology, vol. 42, no. 13, pp. 4639–4647, 2008.


Review Article

Distribution and Fate of Military Explosives andPropellants in Soil: A Review

John Pichtel

Natural Resources and Environmental Management, Ball State University, Muncie, IN 47306, USA

Correspondence should be addressed to John Pichtel, [email protected]

Received 30 November 2011; Revised 26 February 2012; Accepted 19 March 2012

Academic Editor: Jeffrey L. Howard

Copyright © 2012 John Pichtel. This is an open access article distributed under the Creative Commons Attribution License, whichpermits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

Energetic materials comprise both explosives and propellants. When released to the biosphere, energetics are xenobioticcontaminants which pose toxic hazards to ecosystems, humans, and other biota. Soils worldwide are contaminated by energeticmaterials from manufacturing operations; military conflict; military training activities at firing and impact ranges; and openburning/open detonation (OB/OD) of obsolete munitions. Energetic materials undergo varying degrees of chemical andbiochemical transformation depending on the compounds involved and environmental factors. This paper addresses theoccurrence of energetic materials in soils including a discussion of their fates after contact with soil. Emphasis is placedon the explosives 2,4,6-trinitrotoluene (TNT), hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX), and octahydro-1,3,5,7-tetranitro-1,3,5,7-tetrazocine (HMX), and the propellant ingredients nitroglycerin (NG), nitroguanidine (NQ), nitrocellulose (NC), 2,4-dinitrotoluene (2,4-DNT), and perchlorate.

1. Introduction

Energetic compounds, defined as the active chemical com-ponents of explosives and propellants, are necessary both forpeaceful (e.g., demolition and mining) and military pur-poses. Commonly used military energetic compoundsinclude the explosives 2,4,6-trinitrotoluene (TNT), hexa-hydro-1,3,5-trinitro-1,3,5-triazine (RDX), and octahydro-1,3,5,7-tetranitro-1,3,5,7-tetrazocine (HMX) [1]. Nitroglyc-erin (NG), nitroguanidine (NQ), nitrocellulose (NC), 2,4-dinitrotoluene (DNT), and various perchlorate formulationsare employed in missile, rocket, and gun propellants [2,3]. The chemical structures of these compounds appear inFigure 1.

As a result of military activities and due to impropermanagement and disposal practices many energetic sub-stances and their by-products have contaminated environ-ments to levels that threaten the health of humans, livestock,wildlife, and ecosystems. In humans TNT is associated withabnormal liver function and anemia, and both TNT andRDX have been classified as potential human carcinogens[4, 5]. TNT toxicity has been demonstrated using earthworm

reproduction tests [6], and studies with Vibrio fischeri haveestablished TNT as being “very toxic” to aquatic organisms[7]. Mutagenicity studies have been carried out using TNTand its metabolites on Salmonella strains and mammaliancell lines [8–11]. TNT was found to be mutagenic, with somemetabolites more so than the TNT itself.

The effects of RDX on mammals are generally charac-terized by convulsions. Deaths in rats were associated withcongestion in the gastrointestinal tract and lungs [12, 13](oral rat LD50 = 0.07–0.12 g/kg) [14]. Factory employees inEurope and the US have suffered convulsions, unconscious-ness, vertigo, and vomiting after RDX exposure [15]. Infor-mation is limited concerning health effects of HMX [16].The USEPA has established lifetime exposure drinking waterhealth advisory limits for TNT, RDX, and HMX at 2.0, 2.0,and 400 μg/L, respectively [17, 18].

Acute exposure to NG can cause headaches, nausea, con-vulsions, cyanosis, circulatory collapse, or death [19, 20].Chronic exposure may result in severe headaches, halluci-nations, and skin rashes [21]. Perchlorate adversely affectshuman health by interfering with iodine uptake in thethyroid gland [22].


CH3

NO2O2N

NO2

(a)

NO2O2N

NO2

N N

N

(b)

NO2

O2N

NO2

N

NO2NN

N

(c)

O

N+

O−

−O+N

O

OO

O

O

N+O−

(d)

NO2

N

C

NH2H2N

(e)

O−

N+OO

OO

O N+

O−

O

O N+

O

O−O

ON+

N+

N+

OO

O

O−

O−

O

O

OO

O−

(f)

CH3

NO2

NO2

(g)

O

O−Cl

O

O

(h)

Figure 1: Chemical structures of energetic compounds: (a) 2,4,6-trinitrotoluene (TNT); (b) hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX);(c) octahydro-1,3,5,7-tetranitro-1,3,5,7-terazocine (HMX); (d) nitroglycerin (NG); (e) nitroguanidine (NQ); (f) nitrocellulose (NC); (g)2,4-dinitrotoluene (DNT); (h) the perchlorate anion.

Energetic compounds may enter the soil environment vianumerous avenues including [23–28] the following:

(i) ammunition production facilities, for example, wast-ewater lagoons, filtration pits;

(ii) packing or warehouse facilities;

(iii) waste disposal and destruction facilities, for example,open dumps, burn pits, incinerators;

(iv) weapons firing ranges;

(v) weapon impact areas.

Soil contamination by energetics at manufacturing sites,conflict areas, and military ranges is an international con-cern. In the US alone, thousands of military sites are listedas contaminated by energetic compounds [27, 29]. Approx-imately 50 million acres are affected by bombing and othertraining activities [30–32]. An even greater number ofcontaminated sites may exist in Europe and Asia [33].Significant public health emergencies resulting from soilcontamination have launched demands by local citizenry forremediation measures [34]. During the past two decades anincreased environmental awareness has compelled military

agencies in the US, Canada, and many European and Asiannations to identify sites of energetics contamination and toevaluate the impacts of military activities on the quality ofsoil, groundwater, and surface water.

2. Types of Energetic Materials

Energetic compounds are chemicals that, when exposed tophysical or chemical stimuli, decompose extremely rapidlywith the evolution of energy in the form of flame, heat, andlight. In addition, rapid liberation of heat causes the gaseousproducts of the reaction (e.g., N2, CO2, H2O) to expand andgenerate high pressures [33, 35–37].

2.1. Explosives. Explosives are classified as primary, sec-ondary, or tertiary based on their susceptibility to initiation.Primary explosives are highly sensitive to initiation andinclude silver azide, lead styphnate, and mercury fulminate.Primary explosives are often used to initiate secondary explo-sives in a so-called firing train [35]. Common secondaryexplosives include TNT, RDX, HMX (Table 1, Figure 1), and


Table 1: Composition of common military energetic materials (adapted from [36, 38, 39]).

Name Composition

Secondary explosives

Amatex TNT, ammonium nitrate, RDX

Ammonal TNT, ammonium nitrate, aluminium

Anatols TNT, ammonium nitrate

Baratol TNT, barium nitrate

C-4 RDX (91%), plasticizer (9%)

Composition A RDX (91%), wax (9%)

Composition B RDX (60%), TNT (39%), wax (1%)

Cyclotol RDX, TNT

Explosive D Ammonium picrate, picric acid

HTA-3 HMX, TNT, aluminium

Minol TNT, ammonium nitrate, aluminium

Octol HMX (70%–75%), TNT (25%–30%)

Pentolite Ammonium picrate, TNT

Tetrytol Tetryl, TNT

Torpex RDX, TNT, aluminium

Tritonal TNT (80%), aluminium (20%)

Propellants

Single-base smokeless powder (M1; M6; M10) NC, 2,4-DNT; NC, 2,4-DNT; NC, diphenylamine

Double-base smokeless powder (M2, M5, M8) NC, 2,4-DNT; NC, 2,4-DNT; NC, diphenylamine

Triple-base smokeless powder (M30, M31) NC, NG, NQ, ethyl centralite

tetryl (N-methyl-2-4-6-trinitrophenylnitramine). The ener-getic compounds most commonly used in military explosivesinclude TNT, RDX, and HMX. Their environmental fate willbe addressed in this paper. Tertiary explosives, also termedblasting agents, are so insensitive to shock that they cannot bedetonated by reasonable quantities of primary explosive andinstead require a secondary explosive. A common tertiaryexplosive is a physical mixture of ammonium nitrate and fueloil.

Organic secondary explosives can be further divided intonitroaromatics, nitramines, and nitrate esters. Nitroaromat-ics, which include TNT, tetryl, and ammonium picrate, con-tain NO2 groups bonded to carbon atoms on the aromaticring. Nitramines contain NO2 groups bonded to nitrogenpresent within an alicyclic ring, for example, RDX and HMX;nitrate esters contain NO2 groups bonded to an oxygen atomattached to an aliphatic carbon, for example, nitroglycerin.

In addition to the infusion of specialized compounds, anexplosive formulation may contain impurities or decompo-sition by-products. For example, TNT may contain dinitro-toluene and trinitrotoluene isomers [38, 40], and HMX mayoccur as an impurity in RDX [41].

2.1.1. TNT. TNT was first used on a significant scale duringWorld War I. It is one of the most common bulk explosivesin use today both in military ordnance and in mining andquarrying operations. TNT is used as a booster for high-explosive munitions. It is used alone and in mixtures withother energetic compounds (e.g., RDX and HMX) in

explosive formulations including amatol, pentolite, torpex,tritonal, picratol, and others (Table 1) [42].

TNT is chemically and thermally stable, has a low meltingpoint, and is amenable for melt casting [36]. TNT is popularin the military and industry because of its insensitivity toshock and friction, which reduces the risk of accidentaldetonation [35]. The TNT molecule is slightly soluble inwater and has a low vapor pressure and Henry’s law constant(Table 2). The octanol : water partitioning coefficient of TNT(logKow = 1.86) indicates that dissolved TNT will not sorbstrongly to soils and therefore may be mobile in the biosphere[1, 36, 37].

2-amino-4,6-dinitrotoluene (2-A-4,6-DNT) and 4-amino-2,6-dinitrotoluene (4-A-2,6-DNT) are generated inthe biosphere from biotic transformation of TNT nitrogroups to amino groups [43, 44]. Both amino dinitrotolueneisomers are relatively nonvolatile and have solubilities of 17and 36 mg/L, respectively. Amino dinitrotoluenes have lowoctanol:water partitioning coefficients (logKow of 2.8 and2.6); however, they are known to bind covalently to soilorganic and mineral components [1, 36].

2.1.2. RDX. RDX is a highly stable nitramine compound. Itis typically used in mixtures with other explosives [35]. RDXis slightly soluble in water (56.4 mg/L at 25◦C) and has a lowvapor pressure (Table 2). RDX will not readily volatilize fromaqueous solution (Henry’s law constant = 6.3 × 10−8 atm-m3 mol−1) [1] and will not sorb strongly to soil (Kow = 0.90)[36].


2.1.3. HMX. HMX is used as burster charges for artilleryshells [45] and as a component of plastic explosives. HMXhas also been used as an ingredient of solid fuel rocket pro-pellants and to implode plutonium-239 in nuclear weapons[46, 47].

The HMX molecule is of low volatility, has a water sol-ubility of 4.5 mg/L and an octanol : water partitioning coeffi-cient of 0.16 (Table 2) [48]. Dissolved HMX does not readilysorb to soil and therefore may be mobile in the biosphere[3, 37].

2.2. Propellants. Solid propellants for guns, artillery, andmortars comprise low-explosive materials formulated toburn at a controlled rate and produce gases that propelrockets or accelerate projectiles from guns [39, 49]. Theprimary component of gun, artillery, and mortar propellantformulations is commonly a nitro-containing organic chem-ical such as nitrocellulose (NC) often combined with otherenergetic compounds such as nitroglycerin (NG), nitro-guanidine (NQ), or dinitrotoluenes (DNT) [2, 37]. Solidpropellants containing NC are divided into three classesbased on presence of added energetics (Table 1). Single-basepropellants contain NC as the sole energetic material.Double-base propellants contain NC impregnated with anorganic nitrate such as NG. Triple-base propellants includeNC and NG in combination with nitroguanidine (NQ)[36, 49, 50]. Additional ingredients include compoundsthat modify burn rate, binders or plasticizers that facili-tate loading the propellant into the shell, and compoundsthat enhance propellant stability during storage [40].

2.2.1. Nitroglycerin. Nitroglycerin (Figure 1) is a nitrate esterwidely used by the military for the manufacture of pro-pellants and dynamite. Its solubility ranges from 1,250 to1,950 mg/L, and it has a logKow of 1.62 [51, 52]. Nitroglyc-erin was found to be extremely sensitive to slight shocks; toaddress this concern Alfred Nobel in 1866 absorbed nitro-glycerin (75%) in kieselguhr (diatomaceous earth) (25%) tocreate dynamite. Kieselguhr, an inactive ingredient, stabilizesnitroglycerin and makes dynamite a much safer explosive tohandle. Nitroglycerin is often encountered in soils of live-firemilitary training ranges, particularly near firing points [39].

2.2.2. Nitroguanidine. Nitroguanidine (Figure 1) melts at450◦F (232◦C) and decomposes at 480◦F (250◦C). Its sol-ubility at 25◦C is 4.4 g/L, and its vapor pressure is 1.6 ×10−4 bar [53]. Nitroguanidine is not flammable and is anextremely low-sensitivity explosive; however, its detonationvelocity is high. In triple-base smokeless powder NQ reducesthe propellant’s flame temperature without loss of chamberpressure. Nitroguanidine is typically used in large-bore gunswhere barrel erosion and flash are key concerns [39].

2.2.3. Nitrocellulose. Nitrocellulose (Figure 1) is composedof polymerized cellulose chains in which nitrate esters replacemost hydroxyl functions. Other compounds are incorpo-rated to control the physical properties of the propellant, its

burning rate, and long-term stability [39]. Nitrocellulose isinsoluble in water [50].

2.2.4. Dinitrotoluenes. 2,4-Dinitrotoluene (DNT) is used inthe production of smokeless powders, as a plasticizer inrocket propellants and as a gelatinizing and waterproofingagent [54, 55]. Both dinitrotoluene isomers (2,4-DNT and2,6-DNT) may occur as impurities during manufacture ofTNT or may be formed during biotic and abiotic transforma-tion of TNT [3, 19]. 2,4- and 2,6-DNT have similar chemicalproperties. They have low aqueous solubility, are relativelynonvolatile, and have octanol : water partitioning coefficientsof 1.98 and 2.02, respectively [36]. Both are listed as US EPApriority pollutants [56]. 2,4-DNT is often detected in surfacesoils of live-fire military training ranges [57, 58].

2.2.5. Perchlorate. Perchlorate (ClO4−) (Figure 1) is a highly

oxidized (+7) chlorine oxyanion which, when reacted withNH4

+, K+ and Na+, serves as an oxidizer in solid propellantsfor rockets, missiles, explosives, and pyrotechnics [59–62].Composite propellants, used in many rocket motors, typi-cally consist of an organic fuel (such as ammonium picrate)combined with an inorganic oxidizer (commonly perchlo-rate, powdered aluminum, or barium nitrate) and anorganic-binding agent [49, 50]. In a 2001 DOD surveyof weapons systems containing perchlorate, 259 differentmunitions and related items such as fuses, flares, illumina-tion rounds, simulators, and grenades, as well as 41 missilesystems were listed [63].

Most perchlorate salts are highly soluble in water [43];sodium perchlorate has a solubility of about 2 kg/L [64]. Theperiodic replacement and use of solid propellant have result-ed in the discharge of more than 15.9 million kg of per-chlorate salts into the environment [62, 64].

3. Soil Contamination Episodes

Substantial data is available regarding concentrations ofenergetic materials at sites adversely affected by manufac-turing operations [19, 87–91]. Additionally, in recent yearsdetailed information regarding military activities and levelsof energetics residues at firing ranges has become available[2–4, 68, 92] (Figure 2, Table 3).

3.1. Contamination from Manufacturing Operations. Sourcesof soil contamination include explosives machining, castingand curing, improper storage practices, and improper dis-posal of contaminated wastewaters [23, 93]. The PantexPlant (TX) was used during World War II by the US Armyfor loading conventional ammunition shells and bombs.Current operations include development, testing, and fab-rication of high explosive components [90, 94]. The plant’ssolid waste management unit is contaminated with TNT,RDX, and HMX. Concentrations are highest at 10 m depthand continue downward in places to 85 m. Plumes contain-ing explosives have been detected offsite with the potentialfor leaching into the Ogallala aquifer, the region’s primary


Ta

ble

2:P

hysi

cala

nd

chem

ical

prop

erti

esof

sele

cted

expl

osiv

esan

dpr

opel

lan

ts.

En

erge

tic

com

pou

nd

TN

TR

DX

HM

XN

GN

Q2,

4-D

NT

Am

mon

ium

perc

hlo

rate

CA

SN

um

ber

3808

2-89

-200

121-

82-4

0269

14-4

1-0

55-6

3-0

556-

88-7

121-

14-2

1479

7-73

-0C

hem

ical

form

ula

C7H

5N

3O

6C

3H

6N

6O

6C

4H

8N

8O

8C

3H

5(N

O3) 3

CH

4N

4O

C7H

6N

2O

4N

H4C

lO4

MW

227.

1322

2.26

296.

1622

7.00

104.

0718

2.15

117.

49M

elti

ng

poin

t(◦

C)

80–8

220

427

6–28

014

239

64–6

6(d

ecom

pose

s)

Boi

ling

poin

t(◦

C)

240

(exp

lode

s)(d

ecom

pose

s)(d

ecom

pose

s)50

–60

(dec

ompo

ses)

250

300

(dec

ompo

ses)

(dec

ompo

ses)

Solu

bilit

yin

wat

er(m

g/L)

130

(20◦

C)

42(2

0◦C

)5

(25◦

C)

1,25

0to

1,95

0m

g/L

Solu

ble

inw

ater

270

Solu

ble

inco

ldw

ater

,m

eth

anol

,ace

ton

e;2.

49×

105

mg/

LSp

ecifi

cgr

avit

y1.

5–1.

61.

891.

961.

59n

/a∗

1.65

1.95

Vap

orpr

essu

re(1

bar,

20◦ C

)7.

2×

10−9

5.3×

10−1

24.

3×

10−1

72.

6×

10−4

1.6×

10−4

2.2×

10−4

(25◦

C)

Neg

ligib

le

Hen

ry’s

Law

Con

stan

t(b

arm

3m

ol−1

)4.

57×

10−7

to1.

1×

10−8

6.3×

10−8

to1.

96×

10−1

12.

6×

10−1

52.

71×

10−7

atm

m3/m

ole,

25◦ C

(est

.)n

/a1.

86×

10−7

atm

m3/m

ole,

25◦ C

(est

.)N

eglig

ible

Oct

anol

/wat

erpa

rtit

ion

ing

coeffi

cien

t,Lo

gK

ow

1.86

0.86

0.06

11.

62n

/a1.

6–1.

99N

eglig

ible

Ref

11

1,42

43-4

656

,60

1,43

,46–

4849

∗n

/a=

not

avai

labl

e.


Ta

ble

3:Se

lect

edU

San

dC

anad

ian

faci

litie

sco

nta

min

ated

wit

hen

erge

tic

com

pou

nds

.

Site

Stat

eor

prov

ince

,cou

ntr

yE

ner

geti

csde

tect

edC

omm

ents

Ref

.C

old

Lake

Air

Wea

pon

sR

ange

Alb

erta

,Can

ada

TN

T:0

.12–

778

mg/

kg;

31.8

mg/

kg(m

ean

)So

iln

ear

low

-ord

er20

00lb

bom

bcr

ater

.[1

8]

Del

taTo

wer

Alb

erta

,Can

ada

NG

:227

mg/

kg(m

ax.)

Soil.

[18]

Old

anti

tan

kra

nge

Alb

erta

,Can

ada

NG

:256

mg/

kg(m

ean

);95

0m

g/kg

(max

.)So

il,im

pact

ran

ge.

[18]

Are

a8

firi

ng

poin

tA

lber

ta,C

anad

aN

G:1

500

mg/

kgSo

il,10

mbe

hin

dac

tive

firi

ng

poin

t.[1

8]

Dem

olit

ion

ran

geA

lber

ta,C

anad

aR

DX

:11.

4m

g/kg

(mea

n)

HM

X:1

.8m

g/kg

(mea

n)

Soil.

[18]

CFB

Gag

etow

nTr

ain

ing

Are

aN

ewB

run

swic

k,C

anad

a2,

4-D

NT

:491

mg/

kgN

G:t

race

Soil,

burn

pads

.[1

8]

An

tita

nk

ran

geN

ewB

run

swic

k,C

anad

aT

NT

:4m

g/kg

RD

X:<

1m

g/kg

HM

X:6

80m

g/kg

Soil,

impa

ctar

eaof

anti

tan

kra

nge

.[3

]

An

tita

nk

ran

geN

ewB

run

swic

k,C

anad

aN

G:4

700

mg/

kg(i

nfr

ont)

2320

mg/

kg(b

ehin

d)

Soil

infr

ont

offi

rin

glin

eat

anti

tan

kra

nge

.So

ilbe

hin

dfi

rin

glin

eat

anti

tan

kra

nge

.

[3]

New

Bru

nsw

ick,

Can

ada

NG

:656

0m

g/kg

Soil

2m

beh

ind

firi

ng

line

atan

tita

nk

ran

ge.

[65]

Cra

ter

New

Bru

nsw

ick,

Can

ada

TN

T:4

2,20

0m

g/kg

Soil

wit

hin

acr

ater

from

alo

w-o

rder

500

lbbo

mb.

[66]

Cra

ter,

bom

bdr

opN

ewB

run

swic

k,C

anad

aT

NT

:372

0m

g/kg

Soil

2m

from

crat

er.

[3]

CFB

Peta

waw

aO

nta

rio,

Can

ada

TN

T:7

3m

g/kg

RD

X:0

.32

mg/

kgH

MX

:745

mg/

kgN

G:1

45m

g/kg

(max

.)2,

4-D

NT

:53

mg/

kg(m

ax.)

Soil

atan

tita

nk

ran

ge.

[3]

On

tari

o,C

anad

aN

G:2

240

mg/

kg(m

ax.)

Soil

atan

tita

nk

ran

ge.

[18]


Ta

ble

3:C

onti

nu

ed.

Site

Stat

eor

prov

ince

,cou

ntr

yE

ner

geti

csde

tect

edC

omm

ents

Ref

.

CFB

Peta

waw

a,H

otel

Tow

erO

nta

rio,

Can

ada

RD

X:2

.4m

g/kg

(mea

n)

HM

X:4

7m

g/kg

(mea

n)

NG

:44

mg/

kg(m

ean

)2,

6-D

NT

:5.1

mg/

kg(m

ean

)

Soil

atfi

rin

gpo

int.

[18]

CFB

Val

cart

ier

Qu

ebec

,Can

ada

TN

T:2

4m

g/kg

RD

X:4

.6m

g/kg

HM

X:8

03m

g/kg

Soil

atan

tita

nk

ran

ge.

[18]

WA

TC

Wai

nwri

ght

Bri

tish

Col

um

bia,

Can

ada

TN

T:1

26m

g/kg

RD

X:5

.3m

g/kg

HM

X:9

87m

g/kg

Soil.

[3]

WA

TC

Wai

nwri

ght

Bri

tish

Col

um

bia,

Can

ada

TN

T:1

4m

g/kg

(max

.)R

DX

:13.

4m

g/kg

(max

.)H

MX

:0.6

mg/

kg(m

ax.)

NG

:2.7

mg/

kg(m

ax.)

Soil

atde

mol

itio

nra

nge

#14.

[18]

Bri

tish

Col

um

bia,

Can

ada

HM

X:3

,700

mg/

kg(m

ax.)

.So

il.[6

7]Sm

alla

rms

firi

ng

ran

geB

riti

shC

olu

mbi

a,C

anad

aN

G:<

1–52

mg/

kgSo

il.[1

8]

An

tita

nk

targ

ets

Bri

tish

Col

um

bia,

Can

ada

TN

T:3

90m

g/kg

(max

.)So

il.[1

8]H

MX

:161

6m

g/kg

(max

.)N

G:4

,453

mg/

kg(m

ax.)

Don

nel

lyTr

ain

ing

Are

aA

KT

NT

:314

mg/

kgR

DX

:1.4

mg/

kgSo

il,D

elta

Cre

ekIm

pact

Are

a.[6

8]

Fort

Gre

ely

AK

HM

X:0

.11

mg/

kgT

NT

:0.0

6m

g/kg

RD

X:0

.9m

g/kg

HM

X:0

.22

mg/

kg

Soil

ben

eath

81m

mpr

ojec

tile

,low

-ord

erde

ton

atio

n.

[68]

AK

TN

T:1

30m

g/kg

RD

X:3

40m

g/kg

HM

X:4

0m

g/kg

Soil

ben

eath

2.75

inch

rock

et,l

ow-o

rder

deto

nat

ion

.[6

8]

AK

NG

:<0.

01m

g/kg

2,4-

DN

T:4

.3m

g/kg

Soil

atar

tille

ryfi

rin

gpo

int.

[3]

Red

ston

eA

rsen

alA

LPe

rch

lora

te,1

60m

g/L

Gro

un

dwat

er.

[69]


Ta

ble

3:C

onti

nu

ed.

Site

Stat

eor

prov

ince

,cou

ntr

yE

ner

geti

csde

tect

edC

omm

ents

Ref

.R

edst

one

Arm

yG

arri

son

/Mar

shal

lSp

ace

Flig

ht

Cen

ter

AL

RD

X1.

2μ

g/L

Gro

un

dwat

er.S

till

acti

ve,

leas

edto

NA

SA.

[70]

Edw

ards

Air

Forc

eB

ase

CA

Perc

hlo

rate

,160

mg/

LG

rou

ndw

ater

.[6

9]

Fort

Ord

CA

RD

X:0

.5m

g/kg

HM

X:5

87m

g/kg

2,6-

DN

T:0

.9m

g/kg

Soil.

[71]

Nav

alA

irW

eapo

ns

Stat

ion

,Ch

ina

Lake

CA

Perc

hlo

rate

:560

ug/

LG

rou

ndw

ater

.[6

9]

Fort

Car

son

CO

NG

:9.5

–13.

6m

g/kg

2,4-

DN

T:0

.06-

0.27

mg/

kgSo

il.[7

2]

Elg

inA

irFo

rce

Bas

eFL

RD

X:8

.84

mg/

kgH

MX

:0.5

4m

g/kg

Soil

atde

mol

itio

nra

nge

.[4

]

Poh

aku

loa

HI

TN

T:9

.2m

g/kg

RD

X:4

5.6

mg/

kgH

MX

:7.1

mg/

kgN

G:1

.2m

g/kg

Soil

atde

mol

itio

nra

nge

.[3

]

Sch

ofiel

dB

arra

cks

HI

TN

T:3

6m

g/kg

RD

X:5

1m

g/kg

HM

X:9

.1m

g/kg

Soil

ath

and

gren

ade

ran

ge.

[3]

HI

NG

:140

0m

g/kg

Soil

atar

tille

ryra

nge

,0–

10m

beh

ind

firi

ng

poin

t.[2

,73]


Ta

ble

3:C

onti

nu

ed.

Site

Stat

eor

prov

ince

,cou

ntr

yE

ner

geti

csde

tect

edC

omm

ents

Ref

.

Fort

Polk

LAT

NT

:1.2

mg/

kgR

DX

:16

mg/

kgH

MX

:15

mg/

kgSo

il.[7

4]

Cam

pE

dwar

ds,

MM

RM

A

TN

T:<

1–15

mg/

kg;

1.6

mg/

kg(m

ean

)N

G:<

1–15

0m

g/kg

;28

.6m

g/kg

(mea

n)

Perc

hlo

rate

:<1–

351

mg/

kg;

9.6

mg/

kg(m

ean

)

Soil.

[18]

Nav

alSu

rfac

eW

arfa

reC

ente

r,In

dian

Hea

dM

DPe

rch

lora

te,1

,000

ug/

LSu

rfac

ew

ater

.[6

9]

Wel

don

Spri

ng

Ord

nan

ceW

orks

MO

TN

T:6

3,62

1m

g/kg

2,4-

DN

T:1

60m

g/kg

2,6-

DN

T:8

.1m

g/kg

Soil

attw

osi

tes.

Incl

ude

s16

form

erT

NT

and

DN

Tpr

odu

ctio

nlin

es.O

ne

dum

p,si

xbu

rnin

ggr

oun

ds,t

hre

ew

aste

wat

erla

goon

s.

[75,

76]

Fort

Blis

sN

MN

G:0

–0.9

7m

g/kg

Soil.

[77]

NM

TN

T:0

.69

mg/

kgR

DX

:2.1

mg/

kgH

MX

:3.1

mg/

kgSo

il2

mfr

omta

rget

.[1

8]

NM

TN

T:2

520

mg/

kgSo

ilbe

nea

thlo

w-o

rder

155

mm

rou

nd.

[78]

NM

TN

T:1

110

mg/

kgR

DX

:678

mg/

kgH

MX

:149

mg/

kg

Soil

ben

eath

a90

mm

rou

nd.

[78]

Hol

lom

anA

irFo

rce

Bas

eN

MT

NT

:60

mg/

kgSo

ilw

ith

inlo

w-o

rder

bom

bcr

ater

.[3

]

NM

RD

X:1

1.4

mg/

kgH

MX

:1.8

4m

g/kg

Soil

atde

mol

itio

nra

nge

.[3

]

Hol

lom

anA

irFo

rce

Bas

eN

MPe

rch

lora

te16

mg/

LSu

rfac

ew

ater

.[6

9]

Wh

ite

San

dsM

issi

leR

ange

NM

Perc

hlo

rate

21m

g/L

Gro

un

dwat

er.

[69]

US

Arm

yU

mat

illa

Dep

ot,H

erm

isto

n,

Um

atill

aC

oun

tyO

R

TN

T:1

.6m

g/L

(mea

n)

RD

X:1

.0m

g/L

(mea

n)

HM

X:0

.4m

g/L

(mea

n)

2,4-

DN

T:0

.3m

g/L

(mea

n)

Gro

un

dwat

erat

expl

osiv

esw

ash

out

lago

ons.

[79,

80]


Ta

ble

3:C

onti

nu

ed.

Site

Stat

eor

prov

ince

,cou

ntr

yE

ner

geti

csde

tect

edC

omm

ents

Ref

.

Fort

Hoo

dT

XT

NT

:143

,000

mg/

kgR

DX

:323

mg/

kgH

MX

:59

mg/

kg

Soil

ben

eath

low

-ord

er4.

2in

mor

tar.

[72]

TX

TN

T:7

0.2

mg/

kgR

DX

:414

mg/

kgH

MX

:48

mg/

kgSo

ilat

firi

ng

ran

ge.

[18]

Tooe

leA

rmy

Dep

ot(N

orth

Are

a)U

TT

NT

:32,

000

mg/

kgR

DX

:160

0m

g/kg

Surf

ace

soil

at/n

ear

was

hou

tp

onds

.[8

1]

Nan

sem

ond

Ord

nan

ceD

epot

VA

TN

T:1

40,0

00m

g/kg

(198

7)T

NB

:25.

2m

g/kg

2-A

-4,6

-DN

T:

47,0

00m

g/kg

(198

7)

Soil

atT

NT

buri

alsi

te.

Faci

lity

invo

lved

inpr

oces

sin

g,sh

ippi

ng,

rece

ivin

g,an

dde

com

mis

sion

ing

ofor

dnan

ceit

ems

from

1917

to19

60.

[82]

Ban

gor

Nav

alSu

bmar

ine

Bas

eW

AT

NT

:1,5

00m

g/kg

2,4-

DN

T:3

.6m

g/kg

2,6-

DN

T:0

.41

mg/

kgSo

il.[8

3]

Fort

Lew

isW

AT

NT

:15,

100

mg/

kgR

DX<

10m

g/kg

HM

X<

10m

g/kg

Soil

ben

eath

alo

w-o

rder

155

mm

rou

nd.

[84]

WA

NG

:632

mg/

kg(m

ean

)So

il5

to15

mbe

hin

dro

cket

firi

ng

area

.[2

]

Yaki

ma

Trai

nin

gC

ente

rW

AR

DX

:54

mg/

kgH

MX

:5.2

mg/

kgSo

iln

ear

alo

w-o

rder

155

mm

rou

nd.

[78]

WA

NG

:4.6

mg/

kg2,

4-D

NT

:24

mg/

kg2,

6-D

NT

:0.4

mg/

kgSo

ilat

arti

llery

firi

ng

poin

t.[3

]

Form

erD

upo

nt

Cla

ddin

gSi

teat

Cab

inLa

ke(a

kaD

upo

nt

Bar

ksda

leE

xplo

sive

sP

lan

t)

WI

TN

T0.

016

mg/

kg2,

4-D

NT

0.11

mg/

kg2-

A-4

,6-D

NT

0.19

mg/

kg

Soil.

Faci

lity

man

ufa

ctu

red

expl

osiv

esat

the

Bar

ksda

lepl

ant

from

1905

to19

71.

[85]

Cam

pG

uer

nse

yW

YT

NT

:944

0m

g/kg

RD

X:<

10m

g/kg

HM

X:<

10m

g/kg

Soil

ben

eath

aru

ptu

red

500

lbbo

mb.

[78]


Munitionproduction

Wastedisposalfugitive

emissions

StorageMilitary actionor training

Dud(UXO)

SurfaceUXO

BuriedUXO

Blow-in-placeSympatheticdetonation

Long-termcorrosion Long-term

corrosionSympatheticdetonation

Live-firelow order

Live-firehigh order

Figure 2: Possible fates of explosives. (Adapted from [33, 86]).

source of drinking water [33]. The site was placed on theNational Priorities List (NPL) in 1994 [87].

For decades the US military had used unlined evapo-ration/percolation lagoons for disposal of wastewaters frommanufacturing, demilitarization, and load, assemble, andpack (LAP) operations. Many explosives have subsequentlyaccumulated at the surfaces of lagoons, sometimes at con-centrations in the percent range [28, 94]. These areas are asignificant concern relative to long-term soil and groundwa-ter contamination as well as for the potential for accidentaldetonation [18].

Outfalls from explosives manufacturing at Los AlamosNational Laboratory (NM) discharged TNT-, RDX-, andHMX-contaminated waters onto a mesa from 1944 to 1996.Newman et al. [95] detected residues released through burialor dispersion of solid-phase explosives near the Los Alamossite. The majority, however, was released through wastewaterdischarges in surface runoff or into unlined ponds. Explo-sives concentrations in surface soils ranged to more than 20%w/w, and concentrations in surface waters measured up to800 μg/L [95, 96].

From the 1940s through 1977 more than four billionpounds of explosives, primarily TNT and tetryl, were man-ufactured in the MFG Area at the Joliet Army AmmunitionPlant (IL) [88]. The MFG Area contains 139,500 yd3 of soilcontaminated with explosives, primarily TNT, tetryl, andDNT. An additional 13,500 yd3 of soil are enriched withmetals, primarily Pb, and 15,700 yd3 contain both explosivesand metals. A number of groundwater plumes contaminatedwith explosives, volatile organic compounds, and/or metalshave been identified.

In the Joliet Plant LAP Area 12,400 yd3 of soil are contam-inated with TNT, RDX, and HMX, and 17,500 yd3containboth metals and explosives. In addition, several areas contain

unexploded ordnance (UXO). Four separate groundwaterplumes contaminated with explosives have been detected[88].

Explosives processing, handling, and storage took placeat the Savanna Army Depot (IL) until its closure in 1995.Soils are affected by TNT and RDX as well as metals, pesti-cides, and polycyclic aromatic hydrocarbons. TNT and RDX,solvents, and petroleum-related contaminants were detectedin groundwater [97].

An estimated 100,000 tons of TNT were produced at theformer ammunition site WerkTanne in Germany. Duringmanufacturing operations over five million m3 of toxicwastewater were generated. Environmental damage wasexacerbated by the destruction of the facility during Alliedbombing raids in 1944. This site remains highly pollutedtoday with explosives and their metabolites as well as withpolycyclic aromatic hydrocarbons and heavy metals [98].

At the former Explosives Factory Maribyrnong, Victoria,Australia, a 5.5 ton crystalline TNT zone was delineated on aformer waste lagoon. A near-pure TNT layer with an averagethickness of 3 cm was located 10–15 cm below the surface.The TNT contaminant profile in the vadose zone was aresult of leaching, recrystallization, sorption, transformationreactions, precipitation of TNT at the water table interface,and aqueous transport [28].

Large-scale disposal of ammonium perchlorate salts frommanufacturing operations has resulted in contamination ofboth groundwater and surface water, particularly in the west-ern United States. Perchlorate has been detected in watersupplies of 15 million homes in California, Nevada, andArizona [99]. It is also known to contaminate the ColoradoRiver, a major source of irrigation water in the southwesternUS, which could potentially result in perchlorate uptake bycrops [100, 101].


3.2. Contamination from Military Activities

3.2.1. UXOs and Low-Order Detonations. A fired munitionwill experience one of several possible fates. Generally, it willdetonate as intended (a “high-order detonation”). However,it may experience a low-order detonation or not explode atall (i.e., a dud). Unexploded ordnance (UXO) refers to explo-sive, propellant, or chemical-containing munitions that hadbeen armed and fired, yet did not explode due to malfunction[102].

(1) Asia, the Middle East, Africa, and Europe. In 2002 TheViet Nam Ministry of Defense estimated UXO- and land-mine-affected land to comprise “approximately 7-8% ofthe country.” Between 15 and 20% of UXO and minesfrom the war are believed to remain [103]. Official sourcesestimate from 350,000–800,000 tons of war-era ordnance inthe soil [104]. The Ministry of Defense states that “threemillion [antipersonnel] landmines remain in Vietnam’ssoil” [105]. All provinces are affected as well as major cities.Despite extensive clearance operations in the 1990s, land-mines remain a serious problem on the Chinese and Cam-bodian borders. Many UXOs occur on the Laotian border.Minefields remain from the 1954 Dien Bien Phu campaignagainst the French, continuing through border conflicts withChina and the Khmer Rouge in the 1970s [106].

Approximately 85% of all landmines contain TNT [107].Although military grade TNT generally contains about99% of 2,4,6-trinitrotoluene, other components occur,including 2,4-DNT, 2,6-DNT, 1,3-dinitrobenzene, and1,3,5-trinitrobenzene.

More than 580,000 bombing missions were conductedover Laos between 1964 and 1973, with more than 2 mil-lion tons of ordnance dropped. This includes more than 270million cluster submunitions, which are the most commonform of UXO remaining. It is estimated that up to 30% failedto detonate [108]. In Cambodia millions of mortar bombs,rocket-propelled and rifle grenades, artillery shells, clusterbomb submunitions, aircraft bombs, and antipersonnel andantitank mines litter two areas of the country. Only a smallpercentage of land has been cleared of UXO and mines [109].Over 60,000 UXO and mine casualties have been recordedsince 1979 [110].

A total of 1,215 sites contaminated by UXOs have beenidentified in Australia, with the majority in West Australia(334 sites), New South Wales (292), and Queensland (269)[111]. All affected sites comprise military training ranges.

Somaliland, Puntland, and Central and Southern Zone(CSZ) contain hundreds of thousands of UXO includinggrenades, artillery shells, rocket-propelled grenades andmortars. The UXO is amassed at former camps, scattered inthe aftermath of attacks on depots, and deposited duringbattles [112].

Iraq is one of the most energetics-contaminated coun-tries in the world. Affected sites cover an estimated 1,730 km2

and affect 1.6 million people. Contamination includes 20million mines, numerous UXO sites, and many abandonedmunitions sites. Over 50 million cluster bomblets weredropped on Iraqi soil. Landmines are concentrated in Iraqi

Kurdistan, within the major oil infrastructure, and in areasbordering Iran, while UXOs impair areas in the southern andcentral governorates [113].

It is estimated that southern Lebanon was littered withone million undetonated cluster bombs dropped by IsraeliDefense Forces in the last days of the 2006 Israel-Lebanonwar [114, 115]. The Palestinian Occupied Territories areplagued by both UXO and landmines [116]. In 2002 UNICEFconcluded that most minefields dating from the 1967 MiddleEast war were not marked. Israeli military training zones arenot fenced, and UXOs are not collected following trainingexercises. In addition, in most areas of confrontation, Israeliand Palestinian UXOs and improvised explosive devicesremain in the ground [117].

During the war between Armenian forces and Azer-baijan (1988–1994), battle lines frequently shifted, leavingNagorno-Karabakh contaminated with UXO and landmines.The NKR Ministry of Agriculture estimated that 37 mil-lion m2 of arable land and 35 million m2 of pasture areaffected, and 80,000 m2 of vineyards are unusable [118].

Russian officials have admitted to large-scale use ofmines in Chechnya [119]. During 1999-2000 Russian forcesdeployed antipersonnel mines from airplanes, helicopters,and rockets [120, 121]. In 2002 it was estimated that Russiahad planted approximately three million mines during theSecond Chechen War [122]. Throughout 2002 and 2003Chechen rebels used landmines on an almost daily basisagainst both Russian and civilian targets. A true assessmentof the locations and quantities of mines is difficult given thatbattle lines constantly changed combined with factors such asnatural flooding, limited clearance activities, scavenging andreuse, and ongoing fighting [123].

(2) The United States and Canada. Most UXO occurring inthe US and Canada is the result of weapons testing and trooptraining activities. Land affected by UXO includes activemilitary sites, land transferred to private ownership such asFormerly Used Defence Sites, and land no longer used formilitary purposes but still under US Government ownership,such as Base Realignment and Closure sites [124].

Originally 2300 sites in the US and overseas US facilitieswere identified as possible UXO sites. Subsequent investi-gations have narrowed this number to 1400 sites. The totalaffected land area comprises some 10 million acres [125].Some munitions use dates back to the US Civil War andWorld War I [69, 125]. Estimates for remediation of thisacreage have ranged from tens of billions of dollars to morethan $100 billion [124].

Numerous live-fire and demolition ranges have beenstudied at US and Canadian military bases for contaminationby energetics [2, 3, 18, 27, 39, 65–68, 71, 73, 74, 92, 126, 127].Affected sites include antitank rocket, hand grenade, riflegrenade, demolition, tank firing, mortar, artillery, C-130gunship, and bombing ranges. Training is conducted with arange of munitions containing a suite of energetic mixtures.

The primary source of explosive and propellant con-tamination at live-fire ranges is residues from detonation ofmilitary munitions including projectiles (e.g., mortar and


artillery rounds), grenades, landmines, aerial bombs, andmissiles, as well as ordnance demolition charges. Energeticresidues at firing points tend to be composed of propellantformulations, and those at impact areas are high explosivesin warheads [3, 18, 128].

Firing points and targets at live-fire ranges are widelyspaced; therefore, much of the land is uncontaminated byresidues of energetic compounds. A high degree of spatialheterogeneity of residues has been detected [2, 3, 92, 129–132]. Distribution on impact ranges (e.g., from antitank andartillery firing and bomb drops) has been described as ran-domly distributed point sources [18, 133]. Explosives con-centrations spanning five orders of magnitude have beenreported in samples located within 3 m of each other [18,134].

Most detonations during live-fire testing and training arehigh-order (i.e., the round performs as designed). When highorder detonations of artillery rounds, mortars, and handgrenades occur, most of the explosive is consumed and only arelatively small percentage (10−3 to 10−6 percent) of the ini-tial mass is deposited [2, 3, 133–135]. Therefore, high-orderdetonations do not contribute significantly to the overallresidues at firing ranges. As a result of low-order detonationsduring live-fire training large chunks, fibers, slivers, andsoil-size particles (<2 mm) of the original formulations aredispersed over the soil surface [135, 136]. The incidence oflow-order explosions varies among weapon systems.

Taylor et al. [136] estimated that as much as 2% of aTNT-filled 155 mm round remained on the soil surface aftera high-order detonation, which is equivalent to 140 g ofexplosive residues per round. If the round undergoes a low-order detonation, up to 3 kg of TNT would be deposited.Pennington et al. [18] noted a trend of increasing residuemass with decreasing energy of detonation.

At Fort Greely, AK, soil was sampled near a 2.75 inchrocket from a low-order detonation. Concentrations of TNT,RDX and HMX were 130, 340 and 40 mg/kg, respectively[68]. At Fort Bliss (NM), soil near a low-order 155 mm roundcontained 2520 mg/kg TNT [78]. TNT concentrations were143,000 mg/kg near a low-order 4.2 inch mortar [133]. Dur-ing environmental investigations at 23 military firing rangesin the US and Canada, concentrations of TNT and RDX fromComposition B were often hundreds or thousands of mg/kgin soils next to low-order detonations [18]. The bulk chargein US and Canadian army fragmentary munitions is eitherTNT or Composition B. The formula for Composition B isa 60% : 39% ratio of RDX and TNT, containing ∼1% waxas a binder. However, all weapons-grade RDX contains 8to 10% HMX as an impurity. Therefore, the formula forComposition B is closer to 55.2% RDX, 39% TNT, 4.8%HMX, and 1% wax [137].

The most extensively studied US training area is probablyCamp Edwards, Massachusetts Military Reservation (MMR),where military activities have taken place since 1938. Nearly9,500 soil samples and 5,500 groundwater samples havebeen collected and analyzed for contamination by energeticcompounds [8, 18, 138–143]. The energetics most frequentlydetected include, in order of decreasing frequency, perchlo-rate, RDX, HMX, the amino transformation products of

TNT, and 2,4-DNT. The MMR occurs on highly permeablesoils and experiences abundant rainfall. These conditionscreate a “worst case” for contamination of groundwater bymunitions constituents leaching from surface soils [18]. Pri-mary sources of groundwater contamination were low-orderdetonation residues in the impact areas around targets andUXO [27].

3.2.2. Residues as a Function of Weapon System. The typesand concentrations of energetic compounds present in soilare a function of the type of weapon systems used.

Antitank Ranges. The munitions most commonly fired atanti-tank rocket ranges are 66 mm M72 Light AntitankWeapons (also known as Light Anti-Armor Weapon or LAWrockets) and 84 mm AT-4 rockets. The primary charge inLAW rockets is octol, which is composed of HMX and TNTin a 60 : 40 ratio. The double-base M72 propellant contains55% NC, 35% NG, 8% potassium perchlorate, 0.8% ethylcentralite, and 1.2% carbon black [27, 128]. The propellantfor the AT-4 is also double base (NC/NG), but the formula-tion is proprietary [2].

LAW rockets sometimes rupture upon impact withoutdetonating; this is the major source of explosives residuesat impact areas of anti-tank ranges [3]. The primary residuedetected is HMX with concentrations near targets generallyin the hundreds of mg/kg. TNT, RDX, 4-ADNT, and 2-ADNT are often detectable as well, but concentrations areseveral orders of magnitude lower.

High concentrations of HMX were associated with rup-tured LAW rockets at Canadian Force Base (CFB) Valcartier(QC), Western Area Training Center (WATC) Wainwright(BC), CFB Gagetown (NB), and CFB Petawawa (ON), and atUS ranges Fort Ord (CA) and Yakima Training Center (WA)[18, 67, 83, 144]. Concentrations of HMX near tank targetswere as high as 1,640 mg/kg in surface soils at CFB Valcartier[67, 83] and 587 mg/kg at Fort Ord [71]. Soil concentrationsof TNT occurred at levels of about 0.01 that of HMX at bothsites.

Thiboutot et al. [67] studied soil contamination at anti-tank ranges at WATC Wainwright and Canadian ForceAmmunition Depot Dundurn. Relatively high levels of HMXwere detected in surface soils (up to 3,700 mg/kg at Wain-wright), and much lower concentrations of TNT were found.In soil collected near a ruptured M72 rocket at YakimaTraining Center (WA), concentrations of HMX, TNT, andRDX at the 0–0.5 cm depth measured 10,400, 358, and46 mg/kg, respectively. Concentrations declined to 49, 1.7,and 1.5 mg/kg at 6–10 cm depth [3, 78]. At CFB PetawawaHMX and RDX (up to 17 ug/L) were detected in groundwa-ter at impact areas [8].

Nitroglycerin residues are common at the firing pointsof anti-tank rocket ranges due to the use of double-basedpropellant in M72 rockets. Residues have been deposited atdistances up to 100 m in front of the muzzle [3]. The majordeposition of residue, however, is behind the firing line dueto back blast. Studies conducted at anti-tank rocket firingpoints at Yakima Training Center [18, 78], Fort Bliss (TX)


[77], CFB Gagetown [65, 66], CFB Valcartier [131], and CFBPetawawa [145] indicate highest NG concentrations behindthe firing line due to back blast of shoulder-fired rockets.Concentrations at the low percent level are sometimes foundin soil up to 25 m behind the firing line. Nitroglycerin isalso found between the firing line and the target, but theconcentrations are generally several orders of magnitudelower [128].

At firing points at US and Canadian sites, Penningtonet al. [3] found NG to be the primary residue, with soilconcentrations in the hundreds or thousands of mg/kg atdepths from 0 to 25 m. In Fort Lewis (WA), NG concentra-tions in samples 15–25 m, 25–35 m, 35–45 m, and 45–55 mbehind the firing line were 175, 82.4, 13.0, and 3.36 mg/kg,respectively. At a second firing area the mean concentration5–15 m behind the firing line was 936 mg/kg, and 206 mg/kgfor the 15–25 m zone [2]. Nitroglycerin concentrations 5–15 m and 15–25 m behind the firing line at CFB Valcartierwere 788 mg/kg and 339 mg/kg, respectively [131]. At CFBPetawawa 2400 mg/kg was detected 0–10 m behind the firingline, and from 10–20 m behind the line it was 380 mg/kg. Ata second anti-tank rocket range which had been closed for30 years, the mean NG concentration in surface soil was250 mg/kg [145].

At the Gagetown Training Area HMX predominated atthe target, while NG was detected at high levels near thefiring line. The order of energetic residue concentrations wasHMX > NG > TNT > RDX > 2-ADNT and 4-ADNT [65].Nitroglycerin concentrations were as high as 6560 mg/kg at adistance of 2 m behind the firing line [65] and were detectedto a depth of 60 cm. At the front of the firing point at CFBGagetown Pennington et al. [3] measured the majority ofNG in the top 11 cm (range of 6.5–11 mg/kg). Below 11 cmconcentrations were typically <1 mg/kg. HMX, TNT, and NGwere detected in high concentrations in ponds located in thetarget area of the range [18, 74].

2,4-DNT has been detected near firing positions at FortRichardson (AK) and Aberdeen Proving Ground (MD) [146,147]. At the Arnhem anti-tank range (Quebec, Canada),highest perchlorate concentrations were detected in surfacesoils just behind the firing line. Perchlorate was detected inall analyzed groundwater samples [27].

Artillery Ranges. The major munition systems fired intoartillery ranges include 155 mm howitzers, 105 mm how-itzers, 120 mm main tank guns, 81 mm mortars, 60 mm mor-tars, and 120 mm mortars. The explosives used in artilleryand mortar warheads are generally either TNT or Composi-tion B, although some older rounds contained tetryl. Bombsdropped in some ranges contain TNT or tritonal (TNTand aluminum), and 40 mm grenades contain CompositionA5 (RDX). Munitions are delivered using single-, double-,and triple-based gun propellants, and rocket and missilepropellants [3].

Results of soil analyses from several locations indicatethat very low concentrations of explosives residues arewidespread at artillery testing and training ranges. In addi-tion, the distribution of energetics residues is spatially veryheterogeneous [83].

On artillery ranges at Camp Guernsey (WY), Fort Bliss(NM), Fort Hood (TX), Fort Polk (LA), and Fort Greely(AK), TNT, RDX, and HMX concentrations near targetsranged from nondetectable to <1 mg/kg except for low-orderdetonations, where concentrations were three or four ordersof magnitude greater [3, 68, 131, 148]. At artillery andmortar impact areas at Fort Lewis (WA), concentrations ofresidues associated with high-order detonations were typi-cally <1 μg/kg [83]. RDX was detected at <100 μg/kg. Anal-ysis of water samples obtained from monitoring wells andseeps that border the artillery range indicated a low level(<1 μg/L) of RDX contamination. In soil collected near a155 mm round that had undergone a low-order detonationthe TNT concentration was 1.5%. Concentrations were alsohigh in soils collected at 5 and 10 cm depths [83]. At a 105mm howitzer firing point at the Fort Lewis artillery range2,4-DNT was detected at concentrations as high as 237 mg/kgin surface soil [83]. Pennington et al. [3] found thata portion of energetics residues occurred as unburned orpartially burned propellant fibers of lengths ranging from0.4 to 7.5 mm. The unburned fibers contained much higherconcentrations of 2,4-DNT than did partially burned ones.

At gun and mortar firing points at Camp Edwards, MMR,Clausen et al. [138] detected 2,4-DNT in soil, mostly inthe surface to 1 ft depth. Also, 2,6-DNT, diethyl phthalate,n-nitrosodiphenylamine, and di-n-butyl phthalate wereidentified.

Bomb Drops. Air Force ranges have historically measuredhundreds of km2 in size, but current training areas are muchsmaller, generally only tens of hectares. The explosives in USand Canadian Air Force bombs are usually either tritonal(TNT, aluminum powder) or H-6 (TNT, RDX, aluminumpowder). Large-mass high explosive detonations are veryefficient, dispersing only microgram to milligram quantitiesof residue when they detonate high order [135]. As withother ordnance, low-order detonations are the major sourceof residues from bombs [3].

The Cold Lake Air Weapons Range [55, 149] has been inuse for air-to-ground bomb drops for over 40 years. Soil TNTconcentrations at a target ranged from 3 to 408 mg/kg, with amean value of 86 mg/kg over a 50 m radius. Mean concentra-tions of RDX, HMX, 4-ADNT, 2-ADNT, 2,4-DNT, and trini-trobenzene were 0.27, 0.21, 0.71, 1.2, 0.20, and 0.13 mg/kg,respectively [3]. The main sources of RDX and TNT werelow-order detonations and UXO. RDX was the most mobileand persistent contaminant in groundwater, whereas TNTconcentrations were higher but degraded more rapidly.

TNT was detected at the highest concentrations in sur-face and shallow subsurface soil at an impact range at Hol-loman AFB (NM). RDX was generally below detection limitsalthough trace levels were sometimes detected due to use ofC4 to destroy UXO during range maintenance. No evidenceof offsite migration of residues was found with depth or insamples collected along an arroyo that drains the impactrange [18].

At the Donnelly Training Area (AK) [3, 68] residuesincluded TNT (<1–314 mg/kg), RDX (up to 1.4 mg/kg),HMX (<1 mg/kg), 2,4-DNT (<1 mg/kg), and NG


(<1 mg/kg). Only four samples had TNT concentrations>1 mg/kg. One sample collected near a 500 lb bomb partialdetonation had a TNT concentration of 17,300 mg/kg[3, 18].

Hand Grenades. The majority of training at US hand gre-nade ranges is with M67 fragmentation grenades, in whichthe explosive charge is 185 g of Composition B. In Canada,training is generally with C-13 fragmentation grenades; how-ever, they have the same specifications as the M67 [3]. Theprimary energetic residues on grenade ranges are RDX andTNT from Composition B [128, 132].

Variability in soil concentrations of energetics may besubstantial at grenade ranges. At Fort Lewis (WA) and FortRichardson (AK), concentrations of explosives-related com-pounds differed by over two orders of magnitude for samplescollected less than 1 m apart [83].

On grenade ranges at 23 military installations in the USand Canada, RDX concentrations ranged from <1 mg/kg toabout 50 mg/kg [3]. HMX concentrations ranged from<1 mg/kg to 9.1 mg/kg. In several instances, chunks of theexplosive fill were observed near grenade casings, and theirinside surfaces were coated with explosive material.

Jenkins et al. [83] sampled grenade impact areas at FortLewis (WA) and Fort Richardson (AK). An estimated 6,000to 7,000 grenades are thrown on the Fort Lewis range eachyear. RDX was detected in all 96 samples collected at theranges, from both surface and shallow subsurface (depths asgreat as 30 cm at Fort Lewis and 45 cm at Fort Richardson).Median and maximum concentrations of RDX in surfacesoils were 1.6 and 51.2 mg/kg at Fort Lewis and <1 and0.5 mg/kg at Fort Richardson. Pennington et al. [78] detectedTNT and HMX in most soils from grenade ranges, withmaximum concentrations of 40.6 and 5.2 mg/kg, respec-tively. At CFB Petawawa RDX, HMX, and TNT were detectedin relatively low concentrations (<5 mg/kg) [18]. Highestconcentrations were typically measured in the top few cm ofsoil. Mean concentrations were 10.8 and 12.5 times greaterin surface soils than at the 10 cm depth for RDX and HMX,respectively, and about 49 times greater for TNT in thesurface relative to the 10 cm depth [3].

3.3. Demolition and Disposal. Demolition ranges at militaryinstallations are used to destroy munitions that are consid-ered safe to move. Demolition ranges generally measure afew hectares in size [3]. Substantial residues may be dispersedduring demolition events, particularly if they result in low-order detonations or if the C4 does not detonate completely[74]. Some open burning/open detonation (OB/OD) areashave generated kickout, that is, UXO which is propelled fromburn pits when other munitions explode. As a result, someOB/OD areas generate higher concentrations of energeticcompounds than those found in impact areas. OB/OD areaspose an additional concern since many have been usedto dispose of industrial wastes such as paints, solvents,lubricants, and fuels. They may therefore contain a muchwider array of hazardous chemicals than ranges where onlymunitions have burned [151].

RDX is the most common energetic compound at dem-olition ranges, as it is the major component of C4 demolitionexplosive. Nitroglycerin and 2,4-DNT are also frequentlydetected from disposal of excess propellant [2].

RDX and HMX, probably from C4, were detected at ademolition range at Camp Edwards (MMR). At CFB Pet-awawa a demolition range was contaminated with a suite ofexplosives and propellants, primarily RDX and HMX [18].At Elgin AFB (FL) mean concentrations of RDX and HMX insurface soil samples were 8.84 and 0.54 mg/kg, respectively.At Holloman AFB (NM) mean concentrations of RDX andHMX within a 25 m radius around a demolition crater were11.4 and 1.84 mg/kg, respectively [3]. At a site of blow-inplace of 155 mm projectiles at Fort Richardson (AK), averagerecovered residual mass was 14 mg for RDX and 0.84 mg forHMX. No TNT was detected [18, 74]. Soil from the LouisianaArmy Ammunition Plant was contaminated with 10,000,1,900, and 900 mg/kg of TNT, RDX, and HMX, respectively[89]. These levels resulted after incineration of explosives-contaminated soils and sludges. In studies by Jenkins et al.[2], about 1.7% of the original NG in a propellant formula-tion remained after unconfined burning.

4. Environmental Fate of Energetic Materials

Following entry into the terrestrial environment both abioticand biotic processes govern the fate of energetic compounds[36, 37]. The rate and extent of transport and transformationare influenced by the physicochemical properties of thecompounds (e.g., solubility, vapor pressure, Henry’s lawconstant), environmental factors (weather conditions, soilproperties, pH, redox status), and biological factors (popu-lations of energetics-degrading microorganisms). Processesthat influence the environmental fate of explosive com-pounds may be divided into (1) influences on transport(dissolution, volatilization, adsorption) and (2) influenceson transformation (photolysis, hydrolysis, reduction, andbiological degradation) [33]. Figure 3 illustrates the majorfate and transport pathways for energetic materials.

4.1. Dissolution. Energetic residues often occur on the soilsurface as solid particles and chunks resulting from low-order detonation or as partially fragmented UXO. Dissolu-tion in water is the primary mechanism for their transportand dispersion in the biosphere [3, 33]. Once in solution, akey factor affecting fate and transport is advection [86].

Numerous studies have addressed the dissolution mech-anisms of energetic compounds in soil; many, however, haveaddressed dissolution of individual explosives and propellantformulations [152–157]. Results may have limited applica-bility for dissolution of residues on soils at impact zones orfiring ranges because explosives and propellants are typicallyformulated with binders, waxes, stabilizers, and other com-pounds during manufacture. Binders and waxes decrease dis-solution rates of individual explosive compounds [158–160].Dissolution may therefore proceed more slowly than pre-dicted on the basis of solubility of the pure compound [18].

Due to the relatively low aqueous solubility of TNT(130 mg/L), RDX (42 mg/L), and HMX (5 mg/L) (Table 2),


Air phase(volatiles)

Volatilization Deposition

Dispersed/unexplodedordnance

Precipitation Dissolution

Desorption

Adsorption

Precipitation

Dissolution

Biotic uptake

Biotic transformation Bioaccumulation

Transformationproducts

Sorbed phase(humic material, Fe oxides, mineral

components, microorganisms)

Aqueous phase(soil solution, groundwater,

surface water)

Crystallinesolids

Abiotic transformation

Reduction Photolysis Hydrolysis

Figure 3: Suggested fates of energetic compounds in the environment [33, 36, 150].

dissolution of solid particles results in continuous release tothe local environment over extended periods [33, 160]. Instudies performed by Lynch [161] and Lynch et al. [160]using aqueous systems, rate of dissolution followed theorder TNT > HMX > RDX. Dissolution rates at 10◦C wereapproximately 0.0087, 0.0063, and 0.0013 mg/min/cm2 forTNT, HMX, and RDX, respectively [36], and rates doubledwith every 10◦C increase in temperature from 3 to 33◦C[161, 162]. Rates increased as surface area and mixing rateincreased [3, 160, 162–165] and were independent of pH(range 4.2–6.2) [162].

Because the aqueous solubility of HMX is low it tendsto accumulate on the surface while TNT dissolves. DissolvedHMX does not strongly interact with soils and can migratethrough the vadose zone to groundwater. At the FortOrd (CA) anti-tank rocket range HMX was detectable atconcentrations generally <1 mg/kg at depths of 120 cmwhereas TNT, RDX, and amino transformation products ofTNT were not detected below 15 cm [71].

At Fort Greely (AK), Fort Bliss (TX), Fort Lewis (WA),Camp Guernsey (WY), Fort Carson (CO) and elsewhere,HMX and RDX penetrated deeper into the soil profile thandid TNT [3]. This is consistent with the lower soil/waterpartition coefficients for HMX and RDX relative to TNT andthe susceptibility of TNT to attenuation reactions with soilcomponents (see below) [166–168]. RDX and HMX havebeen found in groundwater below several training ranges,but TNT has not.

The behavior of 2,4-DNT and 2,6-DNT in soils has beenevaluated [148, 158, 169]. Dontsova et al. [158] studied thedissolution of 2,4-DNT and 2,6-DNT from propellant for-mulation M1 (87.6% nitrocellulose, 7.3% 2,4-DNT, 0.57%2,6-DNT, 1.06% diphenylamine, 3.48% dibutyl phthalate)

and their subsequent transport in soil. In propellant formu-lations 2,4-DNT is impregnated within an insoluble nitrocel-lulose matrix. M1 dissolution was limited by DNT diffusionfrom the interior of the pellet, resulting in an exponentialdecrease in dissolution rate with time. Dissolution rate washigher for 2,4-DNT than for 2,6-DNT.

At US and Canadian artillery ranges the concentration of2,4-DNT declined from 9.6 mg/kg in the surface 0–3 cm to0.56 mg/kg at 10–20 cm depth [3].

NG is rather mobile in soil in part due to its highsolubility (1,250 to 1,950 mg/L) (Table 2); however, NG andDNT introduced to soil during military training at smallarms ranges did not leach from site soil [170]. The degree towhich NG is available for release is a function of the degreeof deterioration of the nitrocellulose (NC) encapsulation inthe propellant mix [170, 171].

As a consequence of its relatively high solubility and neg-ligible partitioning to soil, perchlorate is not expected topersist in soil for any significant length of time. Once incontact with moisture, solid particulate perchlorate rapidlydissolves and is transported [18].

4.2. Volatilization. At ambient temperatures (approximately0◦–40◦C) most energetic compounds exist as crystallinesolids with vapor pressures ranging from 10−8 to 10−17 atm(Table 2) [36, 37]. Therefore, sublimation (i.e., direct trans-formation from solid to vapor phase) is insignificant.Likewise, few energetic compounds volatilize from theaqueous phase. Compounds with Henry’s law constant(KH) values >10−5 atm-m3/mol may volatilize from aqueoussolutions. With KH values of 10−7 to 10−15 atm-m3/mol,TNT, RDX, HMX, NG, 2,4-DNT, and 2,6-DNT do not


readily volatilize in the aqueous phase [36, 37]. Volatilizationof energetic compounds is therefore a negligible contributionto the biosphere.

4.3. Adsorption. Adsorption refers to a process by which adissolved chemical (solute) accumulates at the surface of aparticle (sorbent). In the current context, surfaces includethose of humic substances, metal oxides and hydroxides, andmicroorganisms. Sorption reactions may include hydropho-bic partitioning, hydrogen bonding, ion exchange, andchemisorption [33, 36, 172, 173]. The extent of partitioningbetween solute and sorbent is a function of their physic-ochemical properties as well as environmental conditions[33, 36].

TNT is reversibly sorbed by soil [167, 173]; hydrogenbonding and ion exchange have been suggested as sorptionmechanisms between the nitro functional groups and soilcolloids [26, 173–175]. Xue et al. [176] determined thatsoil/water partitioning coefficients (Kd) for TNT in surfacesoils ranged from 2.7 to 3.7 L/kg. Pennington and Patrick[173] determined Kd values to range from 2.3 to 11 L/kg. Inaquifer materials, Pennington et al. [177] measured Kdvalues of 0.04–0.27 L/kg.

In general, RDX is sorbed less than TNT by soils [150,167]. Although RDX sorption is minimal, the process isnearly irreversible [174]. Sorption has been described bylinear isotherms [167, 178–180]. Values of Kd for RDX haveranged from 0.21 to 0.33 L/kg [177], 0.12 to 2.37 L/kg [180],and 0.06 to 7.3 L/kg.

Sorption data for HMX is somewhat varied; however,most authors agree that HMX is less sorbed and moremobile than TNT [181]. Values of Kd for HMX have rangedfrom 0–1.6 L/kg (surface soils) [178], 0.09–0.37 L/kg (aquifermaterial) [177], and 0.12–17.7 L/kg (surface soils) [182]. Ingeneral, however, HMX adsorption coefficients are <1 L/kgin aquifer materials and within the range of 1 to 18 L/kg insurface soils [37].

The number of functional groups on nitroaromatic com-pounds like TNT influences their sorption capacity [33].Sheremata et al. [183] determined that Kd values increasedwith an increase in the number of amino groups; that is,sorption of the TNT decomposition product 2,4-diamino-6-nitrotoluene (2,4-DANT) was greater than that for 4-amino-2,6-dinitrotoluene (4-ADNT) (another decomposi-tion product) which was greater than that for TNT [183].

Clay minerals impart a significant effect on sorptionof energetic compounds. Adsorption of TNT onto claysincreased in the order montmorillonite > kaolinite [167]. Ina montmorillonite clay the Kd value for TNT was determinedto be 413 L/kg [184]. Cattaneo et al. [185] determined Kdvalues of 156 and 1.0 L/kg for montmorillonite and kaolinite,respectively. Sorption is additionally influenced by the typesand amounts of exchangeable cations on the clay surface[167, 186, 187]. Haderlein et al. [167] determined adsorptionconstants up to 21,500 L/kg when K+ or NH4

+ were the prin-cipal cations compared to 1.7 L/kg when Al3+, Ca2+, Mg2+, orNa+ was dominant. Brannon et al. [174] demonstrated thatsaturation of cation exchange sites of aquifer solids with K+

and NH4+ resulted in a 9,780% increase in TNT sorption.

In freshwater environments dominated by Ca2+, sorption ofTNT to soil may thus be lower than that observed in a salineenvironment dominated by K+ and Na+. The type of soilor aquifer material and the ionic strength and compositionof the water are therefore critical variables in predictingadsorption [37].

Haderlein et al. [167] determined the Kd for 2,4-DNT reacted with kaolinite, illite, and montmorillonite tobe 690, 3,650, and 740 L/kg, respectively. In contrast, Kdvalues for 2,6-DNT were much lower, with values of 10,52, and 125 L/kg for kaolinite, illite, and montmorillonite,respectively. In adsorption experiments using soils fromCamp Edwards (MMR) the average Kd value for 2,4-DNTwas 3.3 L/kg [169].

Jenkins et al. [2] found that NG and NQ were retainedonly slightly in low organic carbon soils. In studies by Bran-non et al. [188] NQ partition coefficients were very low,ranging from 0.15 to 0.43 L/Kg. Haag et al. [189] reporteda value of <0.1 L/Kg. In soil at small-arms ranges Kd valuesfor 2,4-DNT, 2,6-DNT, and NG ranged from 0.1 to 21.3,0 to 18.2, and 0 to 7.3 L/kg, respectively [190]. Mean Kdvalues were 3.2, 2.6, and 0.9 L/kg, respectively. Soil properties(e.g., organic carbon content, cation exchange capacity, claycontent) imparted a significant effect on Kd. Desorptionexperiments revealed that a portion of these propellants wereirreversibly bound [170]. The lower adsorption coefficientof 2,6-DNT compared to 2,4-DNT in clays is attributed tothe steric hindrance of the NO3 group in the ortho position[158].

Sorption of perchlorate to soil is affected by pH andionic strength. Sorption is negligible in sandy soil under mostconditions, including near-neutral pH [63]. Perchlorate hasbeen observed to adsorb slightly to variable charge soils inlow pH environments, however [191].

The organic carbon fraction in soil plays a significantrole in sorption of energetic compounds. Yamamoto et al.[169] determined that Kd values for TNT, RDX, and 2,4-DNT were dependent on quantity of soil organic carbon.TNT and 2,4-DNT were more strongly adsorbed comparedwith RDX. In studies by Brannon and Pennington [37] andPrice et al. [181], significantly less RDX than TNT wasassociated with soil organic matter. Monteil-Rivera et al.[192] determined that soil organic carbon content did notsignificantly affect HMX sorption. Kd coefficients for HMXwere 2.5 and 0.7 L kg−1 for soil containing 8.4% and 0.33%total organic carbon, respectively.

Numerous investigations of the fate of energetic com-pounds in soils and sediments have used weathered clays,silts, and sands in column and incubation studies [168, 186,189, 193]. The behavior of aged mineral surfaces differsmarkedly from that of newly created surfaces generated byfracturing from detonations, however, Douglas et al. [194]found that fractured soil particles exhibited greater transfor-mation rates for nitroaromatic and nitramine compoundsthan did weathered soil particles. These results may be causedeither by enhanced adsorption to the fractured surfaces or byaccelerated transformation processes.


4.4. Photolysis. Photolysis has been identified as one of themajor processes affecting the transformation of energeticcompounds in waste streams and surface water bodies [175].Alteration of a molecule may occur as a consequence of directabsorption of light energy as influenced by wavelength andintensity or via transfer of energy from a photosensitizedcompound (e.g., peroxide, humic compounds) [36, 195].Photolytic transformation of energetic compounds in soilspresumably occurs only near the soil surface [36].

Phototransformation of TNT results in the formation ofnitrobenzenes, benzaldehydes, azoxydicarboxylic acids, andnitrophenols as a result of the oxidation of methyl groups,reduction of nitro groups and dimer formation [33, 175].Burlinson et al. [196] identified 20 products of TNT pho-tolysis from laboratory irradiation. About 45 to 50 percentof the photodecomposition products of TNT were recoveredin solution with the remainder present as insoluble residues,possibly oligomers of azo and azoxy compounds, which werenot identified [196].

TNT, alone and as a component of Composition B, gen-erated photolysis products more readily than did RDX [18].The rate of photolysis over a 16 h period of irradiation wasrelatively rapid. Photolysis was faster when TNT was mixedwith soil. Results suggest that Composition B photolysis,particularly the TNT component, generates an active mixtureof products occurring on solid surfaces before dissolution,and increasing once in solution [18].

Leachate, groundwater, and surface water of formerammunition sites in Germany were analyzed by Godejohannet al. [197]. Several nitrobenzoic acids were identified;2,4-dinitro-benzoic acid (2,4-DNBA) occurred at 160 μg/kgand 2-amino-4,6-dinitro-benzoic acid (2-A-4,6-DNBA) at86 μg/kg. 4-Amino-nitrobenzoic acid, 2-amino-nitrobenzoicacid, and 2-amino-4-nitrobenzoic acid were detected indrain water samples from a former ammunition plant bySchmidt et al. [198]. Preiss et al. [199] identified severalmethyl-, amino-, and hydroxynitrobenzoic acids in samplesobtained from former German ammunition sites. Henneckeet al. [200] investigated the phototransformation processes ofexplosives in natural water/sediment systems. System pHand the presence of natural photosensitizers influenced TNTphotolysis rate. The major metabolites identified were2-A-4,6-DNBA, 4-A-2,6-DNBA, 2,4,6-trinitrobenzoic acid,1,3,5-trinitrobenzene, and 3,3,5,5-tetranitroazoxybenzene-2,2-dicarboxylic acid.

Photolytic processes may transform RDX with theformation of azoxy compounds, NH3, NO2

−, NO3−, N2O,

formaldehyde, and n-nitroso-methylenediamine [195]. RDXsolution reacted in the presence of sunlight for 28 h untilthe concentration was not distinguishable from background[201]. After treatment, the solution was found not to betoxic to Ceriodaphnia dubia. Additional products of RDXphotolysis included MNX, formate, formamide (CHO-NH2)and urea (CO(NH2)2) [202, 203]. Different products mayform depending on wavelength of light; however, no mecha-nisms have been proposed [204].

Photodegradation of DNTs is possible once dissolved inwater; however, this mechanism is limited in soils because,

similar to TNT [154, 205], dissolution rate is expected to beslow [190].

4.5. Hydrolysis. Hydrolysis involves the reaction of a watermolecule with the functional group of an organic moleculeto form a new carbon-oxygen bond. Amine, amide, nitrile,and carboxylic acid groups are susceptible to hydrolysis.Nitroaromatic compounds, aromatic amines, aldehydes, andbenzenes are generally resistant [36].

Hydrolysis of nitroaromatics such as TNT and aromaticamines occur at strongly elevated pH. Such pH levels typ-ically do not occur in the natural environment; however,alkaline hydrolysis has been investigated as a remediationtechnology for explosives-contaminated soils [33, 36, 206,207]. Hydrolysis of TNT was demonstrated when soil pH was>10 [208, 209]. At pH 12 Bajpai et al. [210] observed >95%reduction in TNT concentration compared to 25% reductionat pH 11. A higher pH was necessary for the destruction of2A-DNT and 4A-DNT compared to TNT [207].

Hydrolysis of RDX and HMX has been noted underalkaline conditions; however, the rate is extremely slow[211, 212]. Balakrishnan et al. [213] observed the formationof NH3, NO2, N2O and formaldehyde following alkalinehydrolysis of RDX and HMX at pH > 10. It was proposedthat initial denitration of cyclic nitramines was sufficient toinduce ring cleavage and spontaneous decomposition of bothcompounds.

Transformation products resulting from hydrolysis reac-tions are similar to those generated during photolysis andbiotransformation, and as a result isolating the effects of thehydrolysis process from other reactions is difficult [36, 175,213].

4.6. Reduction. Energetic compounds containing nitro func-tional groups are susceptible to abiotic reduction, with nitrogroups being reduced to amino groups [33, 37, 175]. Ithas been proposed that abiotic reduction requires activationby solid catalysts such as Fe compounds, clay minerals, ororganic molecules [33, 214].

Reductive transformation of TNT has been widely doc-umented [150, 172, 175, 177, 215–219]. Laboratory studieshave demonstrated the effects of several environmentalfactors on transformation of TNT including pH, redoxstatus, organic carbon levels, cation exchange capacity, andpresence of expandable clays and metallic reducing agents,for example, Fe2+ and Mn2+ [37].

TNT transformations may include reduction of one, two,or all three nitro groups to amines, and coupling of aminotransformation products to form dimers [217]. Formationof the two monoamino products, 2-ADNT and 4-ADNT,are energetically favored [220] and are typically observed inTNT-contaminated soils and groundwater.

TNT reduction is significantly greater under anaerobiccompared with aerobic conditions [193, 219]. TNT dis-appeared from the solution phase of slurry tests underhighly reduced conditions (Eh = −150 mV) following 1 d ofincubation [193]. At Eh values of +500 mV, TNT disappearedafter 4 d.


TNT, RDX, and HMX reduction by Fe (magnetite,ferrous iron, zero valent iron) has been reported[181, 210, 214, 221, 222]. The resulting products (e.g.,aromatic polyamines, hexahydro-1-nitroso-3,5-dinitro-1,3,5-triazine (MNX), hexahydro-1,3-dinitroso-nitro-1,3,5-triazine (DNX), and hexahydro-1,3,5-trinitroso-1,3,5-triazine (TNX)) may be further metabolized via bioticprocesses or may become adsorbed to soil colloids.

The rate of TNT transformation by Fe2+ in the presenceof montmorillonite or kaolinite increased as pH increased[223]. Products were primarily mono amino and azoxycompounds; however, mass balances using radiolabeled TNTindicated the presence of unextractable products. Suppres-sion of the abiotic Fe2+ pathway by addition of ethylenedi-aminetetraacetic acid (EDTA) curbed TNT reduction [223].

Transformation of RDX via formation of mono-, di-,and trinitroso products has been suggested [224, 225] andobserved in soil [226]. Reduction leads to destabilization,ring cleavage, and mineralization of RDX under anaerobicconditions [218, 227, 228]. Gregory et al. [214] observedRDX transformation by Fe2+ in aqueous suspensions of mag-netite. Sequential reduction resulted in the formation of thenitroso intermediates MNX, DNX, and TNX, and NH4

+,N2O, and formaldehyde end-products. Increased pH resultedin greater Fe2+ adsorption which subsequently increasedRDX transformation rate [214]. The nitroso intermediatesof RDX have rarely been observed in field investigations [37].MNX was detected in groundwater at MMR [18].

Several researchers have demonstrated the potential ofzero valent iron (ZVI) for reduction of TNT and RDX in soilor aqueous systems [172, 200, 210, 214, 221, 222, 229, 230].A soil slurry containing 6.4 g RDX/kg was reacted over48 h with 10% ZVI to attain RDX levels within EPArecommended limits [230]. The only product detected wasNH4

+. A study of RDX removal by ZVI in aqueous condi-tions [172] indicated low levels of MNX, DNX, and TNX.These intermediates disappeared after 96 h, and NH+

4 wasgenerated throughout the experiment [204].

Few studies regarding HMX reduction have beenreported. The rate of HMX transformation by ZVI is sig-nificantly less than that of TNT and RDX; however, theapplication of cationic surfactants, which increase HMX sol-ubility, increases transformation rates [229]. Park et al. [229]demonstrated that ZVI-mediated transformation of HMXmay be inhibited in the presence of RDX.

Like many nitrobenzenes [231], 2,4-DNT engages instepwise two-electron transfer reactions with the generationof nitrosobenzene and n-hydroxylaniline intermediates [232]to form aniline.

Although abiotic reduction of energetic compounds isdocumented for soil and aquifer environments, Juhasz andNaidu [36] state that it is practically impossible to distinguishbetween biotic and abiotic transformation.

4.7. Biological Transformation. A variety of microorganismsincluding bacteria and fungi have been demonstrated withthe capacity to degrade both explosive and propellant for-mulations.

4.7.1. TNT. Several TNT biodegradation pathways havebeen reviewed by Crawford [224], Kalderis et al. [33], andJuhasz and Naidu [36] (Figure 4). The TNT molecule servesas a carbon and/or nitrogen source [43]. TNT degradationmay also occur during cometabolism of alternative substrates[33].

Under both aerobic and anaerobic conditions TNT istransformed to the amino derivatives 2-ADNT, 4-ADNT,2,4-DANT, and 2,6-DANT [43, 44]. Triaminotoluene (TAT)may form under anaerobic conditions [33]. Anaerobic min-eralization in the laboratory through TAT to trihydroxytol-uenes, polyphenols, p-cresol, and acetate has been reported[224, 225]. These compounds may be further transformedvia biotic or abiotic processes. Intermediate products maybind to colloidal surfaces, thereby restricting availability forfurther reaction [3].

Fungi mineralize TNT via the actions of nonspecificextracellular enzyme systems (lignin peroxidase, manganeseperoxidase, laccase). A range of fungi acts upon TNT asa function of species, culture conditions, and substrate[33, 36]. Various basidiomycetes (e.g., Agaricus aestivalis,Agrocybe praecox, Clitocybeodora) transform TNT, withdegree of mineralization ranging from 5 to 15%. Severalmicromycetous fungi (e.g., Alternaria sp., Aspergillus terreus,Fusarium sp., Mucor mucedo, Penicillum sp., Rhizoctoniai sp.)are known to transform, but not mineralize, TNT [233].

Phanerochaete chrysosporium, the white rot fungus, is themost widely studied TNT-degrading fungal species. Phan-erochaete chrysosporium was shown to completely trans-form TNT [233]; mineralization ranged from 10 to 40%.The initial reaction involves reduction of nitro groups tonitroso-DNT which may be further transformed to 2-hydroxylamino-4,6-dinitrotoluene and 4-hydroxylamino-2,6-dinitrotoluene (HADNT) and ADNT and DANT. Phane-rochaete chrysosporium and other fungal species may subse-quently transform ADNT and DANT to azo, azoxy, phenolic,and acetylated by-products [33]. Pasti-Grigsby et al. [234]and Spadaro et al. [235] proposed that the azo derivatives areamenable for mineralization.

4.7.2. RDX. Limited aerobic biodegradation of RDX hasbeen observed in soils [236]; however, complete anaerobicbiodegradation has been observed under laboratory condi-tions [157, 211, 237]. RDX is degraded more readily thanTNT, particularly under anaerobic conditions [37, 225, 238–240]. RDX removal from culture was first observed in 1973using purple photosynthetic bacteria [236]; anaerobic photo-synthetic activity was suggested as being responsible for RDXreduction. More recently, anaerobic RDX degradation wasobserved using microbial populations from contaminatedsoil and sewage sludge [225, 241–244] and also under nitrate-reducing [245] and sulfate-reducing [246, 247] conditions.Several cultures required between one week and two monthsto degrade RDX concentrations ranging from 0.015 mMto 0.17 mM. The most rapid degradation of RDX usinganaerobic sludge reported 90% removal of 0.27 mM RDXwithin 2 d [248].


N

O

4A

NHOH

TNT

NHOH4-OHA

2, 4 DA

O2N NO2

CH3 H3C 2, 2 AzNO2NO2

N

CH3

NH2H2N

NO2

2, 6 DA

CH3

NH2

NO2

O2N

2ACH3

O2N

NO2

NH2

CH3

NO2NO2

O2NN

O

N

CH32, 4 AzNO2

O2N

CH3

NO2

CH3

NO2O2N NO2

CH3

O2N

NH2 NH2

NO2

CH3

O2N

NO2

NO2NO2

O2N

H3C CH3 CH3

N

O

N

4, 4 Az

2-OHA

Figure 4: Biological transformation pathway of TNT [217]. Reproduced with kind permission from The American Society for Microbiology.

Degradation under anaerobic conditions may occur viareduction of nitro groups or direct ring cleavage. Sequentialreduction of the nitro groups results in the generation ofMNX, DNX, and TNX [33]. It is proposed that further trans-formation results in the formation of hydroxylamine deriva-tives; however, these products have not been isolated. Hawariet al. [43] proposed that when the first bond in the moleculeis broken, the molecule is destabilized such that it undergoesspontaneous decomposition. The N–N bond is described asthe most likely target for degradation. The products maybe further transformed, ultimately resulting in the produc-tion of nitramine and formaldehyde. Nitramine may beabiotically transformed to N2O. The actions of acetogenicand methanogenic bacteria convert formaldehyde to CO2

[33].Clostridium bifermentans was the first pure strain capable

of anaerobic degradation of RDX to be isolated [240]. Itwas purified from an anaerobic consortium and removed0.23 mM RDX to 25% of its original concentration within24 h. Morganella morganii, which removed 0.33 mM RDXunder oxygen-depleted conditions within 27 d, was the most

efficient isolate from the family Enterobacteriaceae [249].Several strains which could biotransform RDX anaerobicallywere isolated from horse manure, the most effective beingSerratia marcescens which removed 0.23 mM RDX over 10days [250].

In a study by Sheremata and Hawari [251], P. Chrysospo-rium removed RDX (62 mg/L) from liquid medium con-taining glycerol as the main carbon source. Approximately53% of the molecule was mineralized, 11% of the C wasincorporated into fungal biomass, and 28% remained in theaqueous phase as unidentified metabolites. The major by-product (62%) of fungal degradation was N2O.

Bacterial mineralization occurs under aerobic conditionswhere the RDX molecule serves as a nitrogen source(Figure 5) [252, 253]. The first reported (1983) aerobicdegradation of RDX identified three pure strains ofCorynebacterium capable of growing on RDX as a solenitrogen source [254]. The most effective strain removed0.18 mM RDX from culture over 32 h. A consortium ofbacteria from contaminated soil was later reported todegrade 38% of 100 mM RDX over 5 d [255], and a pure


N

N N

Hexahydro-1, 3, 5-trinitro-1, 3, 5-triazine

NON

N N

Mononitrosoderivative

1-hydroxyl-amino-3, 5-dinitro-1, 3, 5-triazine

NHOHN

N N

N-hydroxy-methyl-ethylene-dinitramine

N NH

Methylene-dinitramine

Hydroxymethyl-hydrazine

Methanol

Hydrazine

Nitramide

Formaldehyde

N-hydroxy-methylene

hydrazone

NHOH

HCHO

NH

N

N=ON N

HH

Dimethyl-nitrosamine

1, 1-dimethyl-hydrazine

NHOHN N

HH

NHOHONN N

HH

1-hydroxyl-amino-3-nitroso-5-nitro-1, 3, 5-triazine

NHOH

NON

N N1-hydroxyl-amino-3, 5-dinitroso-1, 3, 5-triazine

NHOHON

NON

N N

NO

NON

N N

Dinitrosoderivative

NOON

NON

N N

Trinitrosoderivative

NHOH

N

N=O

NH

HN

N

O2N NO2

N CH2

CH2OH

H2C CH3

H3C

+

CH2OH

NO2

NO2O2N O2N

O2N NO2 O2N

O2N NO2 O2N

N-hydroxyl-amino-N-nitroso-methylene-

diamine

++

O2N NO2

CH3

CH3H3C

CH2OHH3C CH3

HCHO + H2N − NO2

NH2

NH2

N CH2

HCHO + H2N − NH2

N-hydroxyl-amino-N-nitromethylene-diamine

Figure 5: Biological transformation pathway of RDX [37].

aerobic bacterial strain, also isolated from contaminated soil,removed 0.23 mM RDX from culture over 40 h [256].

4.7.3. HMX. HMX is rather recalcitrant to mineralization[212, 239]. Bacterial degradation may occur via reductionof the nitro groups to form nitroso intermediates (Figure 6)which are subsequently transformed to N2O and formalde-hyde, and eventually to CO2 [36].

Aerobic degradation of HMX by Methylobacterium sp.strain BJ001 was reported by Van Aken et al. [257]. After 55d incubation, 61% of HMX was mineralized to CO2. Axtellet al. [258] reported removal of HMX from contaminatedsoil after amending with nutrients and Pleurotus ostreatus.During 62 d incubation HMX was reduced from 61 ±20 mg/kg to 18 ± 7 mg/kg. Phanerochaete chrysosporiummineralized HMX under nitrogen-limiting conditions [259].After 25 d incubation, 97% of HMX was removed via reduc-tion; 1-NO-HMX accumulated in the culture medium. Theconcentration of 1-NO-HMX peaked after 12 d; a subsequentdecrease in concentration corresponded to release of 14CO2

and the accumulation of 4-nitro-2,4-diazabutanal.

HMX degradation in marine sediment from a militaryUXO disposal site was enhanced in the presence of glucose

[260]. After 50 d the HMX concentration in the aqueousphase (1.2 mg/L) was reduced by 50%. The disappearanceof HMX was accompanied by the formation of a monon-itroso derivative. Boopathy [261] reported degradation tomethanol and chloroform under sulfate-reducing, nitrate-reducing, fermenting, methanogenic, and mixed-electronaccepting conditions using an anaerobic digester sludgeenrichment culture. Rates varied from 22 mg/L on day 0 to<0.05 mg/L on day 11.

4.7.4. Dinitrotoluenes. Extensive studies have been conduc-ted regarding biodegradation of dinitrotoluenes [262, 263].Under aerobic conditions both 2,6-DNT and 2,4-DNTare susceptible to elimination of a nitro group ultima-tely yielding NO2

− and catechols. In anaerobic condi-tions cometabolic reduction of nitro groups to nitroso-,aminonitro-, diamino-, and azoxy compounds has beenobserved [218, 232, 264, 265]. In bioreactor systems degra-dation rates of 86 mg 2,4-DNT/L/h have been determined[264, 266].

4.7.5. NG, NQ, and Perchlorates. A review of degradation ofnitrate esters was provided by Williams and Bruce [267].


O2N

NO2

N

NO2

O2NN

NN

NO2

NNHO CH23

HMX2H

H2O

O O O O

H

NO2

CH2

2H2H

O2N

NO2

NO2

NN

NN

N

NO2

NO2 NO2

O2N

NN

NN

N

N

NN

NO2 NO2

HO CH2 N CH2 N CH2 OH

H2O

NO2

NO2 NO2

NN+ CH2

H2OH2O

OHCH2NHO CH2

HCHO

CO2 + CH2

N2

HCHO

H2O

O O

NN N

N

H H HH

NN

HH

N2O

O

NO

NO NO

N O

Ring cleavage

Anaerobic

Methylenedinitramine

Bis(hydroxymethyl)nitramine

Denitrification

1, 3-H shiftOctahydro-1, 5-dinitroso-3, 7-dinitro-1, 3, 5, 7-tetrazocine

Octahydro-1, 3-dinitroso-5, 7-dinitro-1, 3, 5, 7-tetrazocine

1, 3-H shift

HH

Octahydro-1-nitroso-3, 5, 7-trinitro-1, 3, 5, 7-tetrazocine

Figure 6: Biological transformation pathway of HMX [33]. Reproduced with kind permission from The International Union of Pure andApplied Chemistry.

Several NG-degrading microbial cultures have been iden-tified [268–270]. Biodegradation of nitrate esters occursthrough successive denitrations, each nitro group reactingmore slowly than the previous one (Figure 7) [270–272].Degradation of NG results in the production of glycerol insome cases [269, 272], which is used as a carbon source[2, 273]. Such activity has been observed under both aerobicand anaerobic conditions, using mixed cultures or purestrains including Pseudomonas sp., Agrobacterium radiobacterand Bacillus sp. [269–272, 274, 275]. Enzymes catalyzingthis reaction have been isolated from nitrate ester-degradingorganisms; all have been reductases [273, 276, 277].

Both Pseudomonas putida and P. fluorescens, isolatedfrom NG-contaminated soils, sequentially degraded NG todinitroglycerin (DNG) (C3H6NO7) and glycerol mononi-trate (GMN) (C3H7NO5) isomers [278]. Wendt et al. [270]showed that without a carbon supplement the ability ofPseudomonas to degrade NG is significantly reduced. Awastewater disposal lagoon at a former NG production plantcontained Arthrobacter ureafaciens, Klebsiella oxytoca, and aRhodococcus [279]. These bacteria were capable of degrading

NG and producing GDN and GMN. Moreover, Rhodococcusspecies removed the final nitro group from GMN thusachieving complete NG biodegradation.

Penicillium corylophilum Dierckx, Bacillus thuringiensis,and B. cereus were capable of denitrating NG to glycerol[269, 279]. Phanerochaete chrysosporium denitrified NG in anaerobic environment without added carbon. However, thisprocess was greatly enhanced when a carbon source wasadded. Christodoulates et al. [272] showed that mixedbacterial cultures in anaerobic microcosms mineralized NGmore rapidly in the presence of a carbon source. Nitroglyc-erin mineralization was complete in 26 d with addition of2000 mg/L glucose as compared with 114 days without acarbon source [280].

In a study by Brannon et al. [188], concentrations of NQin solution phase did not change over time, which reflectsthe lack of susceptibility of NQ to aerobic biodegradation inactivated sludge.

Degradation of perchlorate is carried out directly bysoil microorganisms and also by enzymes such as perchlo-rate reductase and chlorite dismutase [281]. A number of


O2N NO2 NO2

O O O

H2C CH

CH2

NGOH

NO2O2N

OO

H2C CH

CH2

NO2

O OH

H2C CH

CH2

HO

OH OH OH

H2C CH2CH

CH

O OH OH

H2C CH2

NO2

CH2H2C CH

OH OO

NO2 NO2

Figure 7: Biological transformation pathway for NG [272]. Repro-duced with kind permission from Elsevier.

perchlorate-reducing bacteria have been isolated; they areall Gram-negative, facultative anaerobes [64]. Tipton et al.[101], using surface and subsurface soils and dredge tail-ings, demonstrated that perchlorate biodegradation requiresanaerobic conditions, an adequate carbon source, and anactive perchlorate-degrading microbial population. Brannonet al. [188] observed a slight decrease in solution phaseperchlorate concentration in clay at −150 mV and pH 7.0.

Perchlorate degradation is generally accepted to proceedvia the pathway, ClO4

− → ClO3− → ClO2

− → Cl− + O2

(Figure 8) [282]. A wide variety of organic compoundsincluding alcohols and carboxylic acids are used as growthsubstrates by perchlorate-reducing bacteria although theiruse is strain-dependent [64]. Acetate-oxidizing chlorate-reducing bacteria represent a significant population fromvaried sources, including noncontaminated soils, hydro-carbon-contaminated soils, aquatic sediments, papermillsludges, and livestock waste lagoons [283]. Acetate has beenmost frequently used as a substrate for heterotrophic per-chlorate reduction [284–288], although hydrogen or formateis required as an electron donor in limited cases [285]. Per-chlorate reduction by Citrobacter [289] was sustained onacetate, but yeast extract improved growth.

5. Summary and Conclusions

Energetic materials have contaminated soils worldwide frommanufacturing operations, military conflict, and militarytraining activities. Explosives such as TNT, RDX, and HMXand propellants such as NG and perchlorates present thegreatest concern to public health and the environmentbecause they are manufactured and used in the greatestquantities. Types of residues, concentrations, and distribu-tions differ depending on the type of range (e.g., bombdrops versus demolition) and munition used. Distribution ofenergetics in soil is highly variable—concentrations vary by

AcetatePerchloratereductase

Acetate

Acetate

CO2 + H2OCl− + O2

CO2, H2O, biomass

CO2, H2O, biomass

ClO4−

ClO3−

ClO2−

Figure 8: Biological transformation pathway for perchlorate(Adapted from [64, 282]).

orders of magnitude across short distances on the surface,and often with only slight depth. Both low-order detonationsand UXO comprise the highest concentrations of energeticscontamination.

Energetic compounds undergo varying degrees of chem-ical and biochemical transformation, depending on thecompounds involved and environmental factors. Both RDXand perchlorate appear to be common groundwater contam-inants; however, several energetics, including some decom-position products (e.g., DNTs, amino-DNTs), pose healthand environmental hazards.

In several countries assessment of soil contamination byexplosives residues and UXOs has been highly effective,resulting in extensive cleanup; however, many nations con-tinue to be plagued by massive dispersion of UXOs andmunitions, often from conflicts decades past. There is anurgent need to identify, limit public access to, and removesuch hazards to ensure public safety, promote local economicdevelopment, and protect local natural resources over thelong term.

References

[1] S. Thiboutot, G. Ampleman, and A. Hewitt, “Guide forcharacterization of sites contaminated with energetic materi-als,” Tech. Rep. ERDC/CRREL TR-02-1, US Army EngineerResearch and Development Center, Hanover, NH, USA,2002.

[2] T. F. Jenkins, J. C. Pennington, G. Ampleman et al., “Char-acterization and fate of gun and rocket propellant residueson testing and training ranges: interim report 1,” Tech.Rep. ERDC TR-07-01, Strategic Environmental Research andDevelopment Program, Vicksburg, Miss, USA, 2007.

[3] J. C. Pennington, T. F. Jenkins, G. Ampleman et al., “Dis-tribution and fate of energetics on DOD test and trainingranges: final report,” ERDC TR-06-13, US Army Corps


of Engineers Engineer Research and Development Center,Vicksburg, Miss, USA, 2006.

[4] ATSDR, “2,4,6-Trinitrotoluene (TNT) fact sheet,” 1996,http://www.atsdr.cdc.gov/tfacts78.html.

[5] ATSDR, “RDX fact sheet,” 1996, http://www.atsdr.cdc.gov/tfacts78.html.

[6] P. Y. Robidoux, C. Svendsen, J. Caumartin et al., “Chronictoxicity of energetic compounds in soil determined using theearthworm (Eisenia andrei) reproduction test,” Environmen-tal Toxicology and Chemistry, vol. 19, no. 7, pp. 1764–1773,2000.

[7] O. Drzyzga, T. Gorontzy, A. Schmidt, and K. H. Blotevogel,“Toxicity of explosives and related compounds to the lumi-nescent bacterium Vibrio fischeri NRRL-B-11177,” Archives ofEnvironmental Contamination and Toxicology, vol. 28, no. 2,pp. 229–235, 1995.

[8] S. E. George, G. Huggins-Clark, and L. R. Brooks, “Use ofa Salmonella microsuspension bioassay to detect the muta-genicity of munitions compounds at low concentrations,”Mutation Research, vol. 490, no. 1, pp. 45–56, 2001.

[9] B. Lachance, P. Y. Robidoux, J. Hawari, G. Ampleman, S.Thiboutot, and G. I. Sunahara, “Cytotoxic and genotoxiceffects of energetic compounds on bacterial and mammaliancells in vitro,” Mutation Research, vol. 444, no. 1, pp. 25–39,1999.

[10] E. L. Tan, C. H. Ho, W. H. Griest, and R. L. Tyndall, “Muta-genicity of trinitrotoluene and its metabolites formed dur-ing composting,” Journal of Toxicology and EnvironmentalHealth, vol. 36, no. 3, pp. 165–175, 1992.

[11] W.-Z. Whong and G. S. Edwards, “Genotoxic activityof nitroaromatic explosives and related compounds inSalmonella typhimurium,” Mutation Research, vol. 136, no. 3,pp. 209–215, 1984.

[12] L. J. Burdette, L. L. Cook, and R. S. Dyer, “Convulsantproperties of cyclotrimethylenetrinitramine (RDX): sponta-neous, audiogenic, and amygdaloid kindled seizure activity,”Toxicology and Applied Pharmacology, vol. 92, no. 3, pp. 436–444, 1988.

[13] W. F. von Oettingen, D. D. Donahue, H. Y. Yagonda, A. R.Monaco, and M. R. Harris, “Toxicity and potential dangers ofcyclotrimethylenetrinitramine (RDX),” Journal of IndustrialHygiene and Toxicology, vol. 31, pp. 21–31, 1949.

[14] C. Smith-Simon and S. Goldhaber, “RDX - ATSDR toxi-cological profile,” Report 205-93-0606, US Department ofHealth and Human Services, Atlanta, Ga, USA, 1999.

[15] A. S. Kaplan, C. F. Berghout, and A. Peczenik, “Humanintoxication from RDX,” Archives of Environmental Health,vol. 10, pp. 877–883, 1965.

[16] ATSDR, “HMX fact sheet,” 1996, http://www.atsdr.cdc.gov/tfacts98.html.

[17] A. Crockett, H. Craig, and T. Jenkins, “Field sampling andselecting on-site analytical methods for explosives in water,”Report EPA/600/S-99/002, US Environmental ProtectionAgency, Washington, DC, USA, 1999.

[18] J. C. Pennington, T. F. Jenkins, G. Ampleman et al., “Dis-tribution and fate of energetics on DOD test and trainingranges: interim report 6,” TR 06-12, Strategic Environmen-tal Research and Development Program, US Army Corpsof Engineers Engineer Research and Development Center,Vicksburg, Miss, USA, 2006.

[19] M. Sittig, Handbook of Toxic and Hazardous Chemicals, NoyesPublications, Park Ridge, NJ, USA, 3rd edition, 1991.

[20] W. N. Rom, Environmental and Occupational Medicine, Little,Brown and Company, Boston, Mass, USA, 2nd edition, 1992.

[21] OSHA, “Occupational safety and health guideline fornitroglycerin,” http://www.osha.gov/SLTC/healthguidelines/nitroglycerin/recognition.html.

[22] ATSDR, “Toxic substances portal—perchlorates,” 2011,http://www.atsdr.cdc.gov/toxfaqs/.

[23] E. P. H. Best, S. L. Sprecher, S. L. Larson, H. L. Fredrickson,and D. F. Bader, “Environmental behavior of explosivesin groundwater from the Milan army ammunition plantin aquatic and wetland plant treatments. Removal, massbalances and fate in groundwater of TNT and RDX,”Chemosphere, vol. 38, no. 14, pp. 3383–3396, 1999.

[24] H. R. Beller and K. Tiemeyer, “Use of liquid chromatogra-phy/tandem mass spectrometry to detect distinctive indica-tors of in situ RDX transformation in contaminated groundwater,” Environmental Science and Technology, vol. 36, pp.2060–2066, 2002.

[25] R. F. Spalding and J. W. Fulton, “Groundwater munitionresidues and nitrate near Grand Island, Nebraska, U.S.A.,”Journal of Contaminant Hydrology, vol. 2, no. 2, pp. 139–153,1988.

[26] K. Spiegel, J. V. Headley, K. M. Peru, N. Haidar, andN. P. Gurprasard, “Residues of explosives in groundwaterleached from soils at a military site in Eastern Germany,”Communications in Soil Science and Plant Analysis, vol. 36,no. 1–3, pp. 133–153, 2005.

[27] R. Martel, M. Mailloux, U. Gabriel, R. Lefebvre, S. Thiboutot,and G. Ampleman, “Behavior of energetic materials inground water at an anti-tank range,” Journal of EnvironmentalQuality, vol. 38, no. 1, pp. 75–92, 2009.

[28] R. Martel, T. J. Robertson, D. M. Quan et al., “2,4,6-Trinitrotoluene in soil and groundwater under a waste lagoonat the former Explosives Factory Maribyrnong (EFM),Victoria, Australia,” Environmental Geology, vol. 53, no. 6, pp.1249–1259, 2008.

[29] S. Hains, R. Martel, R. Lefebvre et al., in Proceedings of the54th Canadian Geotechnical Conference: an Earth Odyssey, pp.16–19, Calgary, Canada, September 2001.

[30] J. Clausen, J. Robb, D. Curry, M. Wojtas, and B. Gallagher,“Analytes of interest at military ranges,” in Proceedings ofthe National Defense Industry Association Annual Meeting,Arlington, Va, USA, April 2003.

[31] D. Armstrong, “US presence on foreign soil is tainted,” BostonGlobe, November 15, 1999.

[32] D. Armstrong, “More costly cleanup on horizon,” BostonGlobe, November 14, 1999.

[33] D. Kalderis, A. L. Juhasz, R. Boopathy, and S. Comfort,“Soils contaminated with explosives: environmental fate andevaluation of state-of-the-art remediation processes (IUPACtechnical report),” Pure and Applied Chemistry, vol. 83, no. 7,pp. 1407–1484, 2011.

[34] J. Eriksson, S. Frankki, A. Shchukarev, and U. Skyllberg,“Binding of 2,4,6-trinitrotoluene, aniline, and nitrobenzeneto dissolved and particulate soil organic matter,” Environ-mental Science and Technology, vol. 38, no. 11, pp. 3074–3080,2004.

[35] J. Pichtel, Terrorism and WMDs: Awareness and Response,CRC Press, Boca Raton, Fla, USA, 2011.

[36] A. L. Juhasz and R. Naidu, “Explosives: fate, dynamics, andecological impact in terrestrial and marine environments,”Reviews of Environmental Contamination and Toxicology, vol.191, pp. 163–215, 2007.


[37] J. M. Brannon and J. C. Pennington, “Environmental fateand transport process descriptors for explosives,” ERDC/ELTR-02-10, U.S Army Engineer Research and DevelopmentCenter, Vicksburg, Miss, USA, 2002.

[38] D. C. Leggett, T. F. Jenkins, and R. P. Murrmann, “Compo-sition of vapors evolved from military TNT as influencedby temperature, solid composition, age and source,” SpecialReport 77-16, Cold Regions Research and EngineeringLaboratory, Hanover, NH, USA, 1977.

[39] T. F. Jenkins, G. Ampleman, S. Thiboutot et al., “Charac-terization and fate of gun and rocket propellant residues ontesting and training ranges: final report,” TR 08-01, StrategicEnvironmental Research and Development Program, ColdRegions Research and Engineering Laboratory, US ArmyEngineer Research and Development Center, Hanover, NH,USA, 2008.

[40] M. E. Walsh, T. F. Jenkins, P. S. Schnitker, J. W. Elwell,and M. H. Stutz, “Evaluation of SW846 Method 8330 forcharacterization of sites contaminated with residues of highexplosives,” CRREL Report 93-5, US Cold Regions Researchand Engineering Laboratory, Hanover, NH, USA, 1993.

[41] US EPA, Contaminated Site Clean-Up Information, Technol-ogy Innovation and Field Services Division, US Environmen-tal Protection Agency, Washington, DC, USA, 2010.

[42] Global Security, “Explosives—nitroaromatics,” 2008, http://www.globalsecurity.org/military/systems/munitions/explo-sives-nitroaromatics.htm.

[43] J. Hawari, “Biodegradation of RDX and HMX: from basicresearch to field application,” in Biodegradation of Nitroaro-matic Compounds and Explosives, pp. 277–310, Plenum Press,New York, NY, USA, 2000.

[44] H. Ullah, A. A. Shah, F. Hasan, and A. Hameed, “Biodegra-dation of trinitrotoluene by immobilized Bacillus Sp. Yre1,”Pakistan Journal of Botany, vol. 42, no. 5, pp. 3357–3367,2010.

[45] J. Yinon, Toxicity and Metabolism of Explosives, CRC Press,Boca Raton, Fla, USA, 1990.

[46] J. H. Sullivan, H. D. Puttman, M. A. Keirn, B. C. Pruitt, J.C. Nichols, and J. T. McClave, A Summary and Evaluationof Aquatic Environmental Data in Relation to EstablishingWater Quality Criteria for Munitions Unique Compounds:Part 2. Nitroglycerin, Water and Air Research, US ArmyMedical Research and Development Command, Gainesville,Fla, USA, 1979.

[47] R. Boopathy, “Enhanced biodegradation of cyclotetrameth-ylenetetranitramine (HMX) under mixed electron-acceptorcondition,” Bioresource Technology, vol. 76, pp. 241–244,2001.

[48] F. Monteil-Rivera, L. Paquet, S. Deschamps, V. K. Balakrish-nan, C. Beaulieu, and J. Hawari, “Physico-chemical measure-ments of CL-20 for environmental applications: comparisonwith RDX and HMX,” Journal of Chromatography, vol. 1025,pp. 125–132, 2004.

[49] M. R. Walsh, M. E. Walsh, A. D. Hewitt, and C. M. Collins,“Field-expedient disposal of excess artillery propellants,” inProceedings of the SERDP & ESTCP’s Partners in Environmen-tal Technology Technical Symposium & Workshop, Washingto,DC, USA, December 2008.

[50] National Forensic Science Technology Center, “Propellants,”Firearm Examiner Training, 2011, http://projects.nfstc.org/firearms/module05/fir m05 t04 01.htm.

[51] D. H. Rosenblatt, E. P. Burrows, W. R. Mitchell, and D. L.Palmer, “Organic explosives and related compounds,” in The

Handbook of Environmental Chemistry , O. Hutziner, Ed., vol.3, pp. 195–234, Springer, Berlin, Germany, 1989.

[52] D. C. Leggett, T. F. Jenkins, and R. P. Murrmann, “Compo-sition of vapors evolved from military TNT as influencedby temperature, solid composition, age and source,” SpecialReport 77-16, Cold Regions Research and EngineeringLaboratory, Hanover, NH, USA, 1977.

[53] Fisher Scientific, “Material Safety Data Sheet, Nitroguani-dine, 99%, Moistened With ca 25% Water,” 2008, https://fscimage.fishersci.com/msds/07334.htm.

[54] G. Bordeleau, R. Martel, G. Ampleman, and S. Thiboutot,“Environmental impacts of training activities at an airweapons range,” Journal of Environmental Quality, vol. 37,pp. 308–317, 2008.

[55] ATSDR, Toxicological Profile for 2,4-Dinitrotoluene and 2,6-Dinitrotoluene, US Public Health Service, US Department ofHealth and Human Services, Atlanta, Ga, USA, 1989.

[56] Code of Federal Regulations (CFR), Appendix A to Part423-126 Priority Pollutants, Title 40-Protection ofEnvironment, http://ecfr.gpoaccess.gov/cgi/t/text/text-idx?c=ecfr&tpl=%2Findex.tpl.

[57] K. A. Thorn, J. C. Pennington, K. R. Kennedy, L. G. Cox, C. A.Hayes, and B. E. Porter, “N-15 NMR study of the immobi-lization of 2,4- and 2,6-dinitrotoluene in aerobic compost,”Environmental Science and Technology, vol. 42, pp. 2542–2550, 2008.

[58] M. Uchimiya, “Reductive transformation of 2,4-dinitrotol-uene: roles of iron and natural organic matter,” AquaticGeochemistry, vol. 16, pp. 547–562, 2010.

[59] E. T. Urbansky, “Perchlorate chemistry: implications foranalysis and remediation,” Bioremediation, vol. 2, pp. 81–95,1998.

[60] R. Q. Gullick, M. W. Lechvallier, and T. A. S. Barhorst,“Occurrence of perchlorate in drinking water sources,” Jour-nal of the American Water Works Association, vol. 39, pp. 66–77, 2001.

[61] B. E. Logan, “Assessing the outlook for perchlorate reme-diation,” Environmental Science and Technology, vol. 35, pp.482A–487A, 2001.

[62] W. E. Motzer, “Perchlorate: problems, detection, and solu-tions,” Environmental Forensics, vol. 2, pp. 301–311, 2001.

[63] C. W. Trumpolt, M. Crain, C. D. Cullison, S. J. P. Flanagan,L. Siegel, and S. Lathrop, “Perchlorate: sources, uses, andoccurrences in the environment,” Remediation Journal, vol.16, no. 1, pp. 65–89, 2005.

[64] J. Xu, Y. Song, B. Min, L. Steinberg, and B. E. Logan, “Micro-bial degradation of perchlorate: principles and applications,”Environmental Engineering Science, vol. 20, no. 5, pp. 405–422, 2003.

[65] S. Thiboutot, G. Ampleman, A. Marois et al., “Environmentalconditions of surface soils, CFB gagetown training area:delineation of the presence of munitions-related residues(phase III, final report),” Tech. Rep. DREV-TR-2004-205,Defence Research and Development Canada-Valcartier,Quebec, Canada, 2004, http://cradpdf.drdc.gc.ca/PDFS/unc57/p522641.pdf.

[66] S. Thiboutot, G. Ampleman, A. Marois et al., “Environmentalcondition of surface soils and biomass prevailing in thetraining area at CFB Gagetown, New Brunswick,” DRDCValcartier TR 2003-152, Defence Research and DevelopmentCanada-Valcartier, Quebec, Canada, 2004.

[67] S. Thiboutot, G. Ampleman, A. Gagnon et al., “Character-ization of antitank firing ranges at CFB valcartier, WATC


Wainwright and CFAD Dundurn,” Quebec Report DREVR-9809, Defence Research Establishment Valcartier, Quebec,Canada, 1998.

[68] M. E. Walsh, C. H. Racine, T. F. Jenkins, A. Gelvin, and T.A. Ranney, “Sampling for explosives residues at Fort Greely,Alaska,” ERDC/CRREL TR-01-15, US Army Cold RegionsResearch and Engineering Laboratory, Hanover, NH, USA,2001.

[69] Government Accountability Office (GAO), “DOD opera-tional ranges—more reliable cleanup cost estimates and aproactive approach to identifying contamination areneeded,” GAO-04-601, 2004, http://www.gao.gov/htext/d04601.html.

[70] ATSDR, “Redstone Army Garrison/Marshall Space FlightCenter Huntsville, Madison County, Alabama,” PublicHealth Assessments and Health Consultations, 2005 http://www.atsdr.cdc.gov/HAC/pha/RedstoneArmy/RedstoneArmy-PHA071205.pdf.

[71] T. F. Jenkins, M. E. Walsh, P. G. Thorne et al., “Site char-acterization at the inland firing range impact area at FortOrd,” Special Report 98-9, US Army Cold Regions Researchand Engineering Laboratory, 1998.

[72] A. D. Hewitt, T. F. Jenkins, M. E. Walsh, M. R. Walsh, and S.Taylor, “RDX and TNT residues from live-fire and blow-in-place detonations,” Chemosphere, vol. 61, pp. 888–894, 2005.

[73] A. D. Hewitt, T. F. Jenkins, T. A. Ranney, D. Lambert, andN. Perron, “Characterization of energetic residues at firingranges: schofield barracks and training area,” Distributionand Fate of Energetics on DoD Test and Training Ranges,Interim Report 4 ERDC TR-04-4, US Army EngineerResearch and Development Center, Vicksburg, Miss, USA,2004.

[74] J. C. Pennington, T. F. Jenkins, S. Thiboutot et al., “Distribu-tion and fate of energetics on DoD test and training ranges:interim report 5,” TR 05-2, Strategic Environmental Researchand Development Program, US Army Corps of EngineersEngineer Research and Development Center, Vicksburg,Miss, USA, 2005.

[75] ATSDR, Weldon Spring Ordnance Works, Weldon Spring,St. Charles County, Missouri, Public Health Assessmentsand Health Consultations, 2010, http://www.atsdr.cdc.gov/HAC/pha/pha.asp?docid=871&pg=2.

[76] Environmental Science & Engineering, Draft Final RemedialInvestigation/Baseline Risk Assessment Explosives-MunitionsManufacturing Areas Operable Unit, vol. I-V, Crab OrchardNational Wildlife Refuge, Louis, Mo, USA, 1994.

[77] USACHPPM, “Training range site characterization and riskscreening, regional range study, Dona Ana Range, Fort Bliss,Texas, May 2002,” Geohydrologic Study 38-EH-6807-02, USArmy Center for Health Promotion and Preventive Medicine,Aberdeen Proving Ground, MD, USA, 2004.

[78] J. C. Pennington, T. F. Jenkins, G. Ampleman et al., “Dis-tribution and fate of energetics on dod test and trainingranges: interim report 2,” ERDC TR-02-8, US Army EngineerResearch and Development Center, Vicksburg, Miss, USA,2002.

[79] US Environmental Protection Agency, “EPA superfundrecord of decision: Umatilla Army Depot (Lagoons), Hermis-ton, OR,” EPA/ROD/R10-94/094, 1994.

[80] ATSDR, US Army Umatilla Depot Activity. Hermiston,Umatilla County, Oregon, Public Health Assessments andHealth Consultations, 2009, http://www.atsdr.cdc.gov/HAC/pha/PHA.asp?docid=290&pg=1.

[81] ATSDR, Tooele Army Depot (North Area), Tooele, TooeleCounty, Utah, Public Health Assessments and HealthConsultations, 2010, http://www.atsdr.cdc.gov/HAC/pha/pha.asp?docid=108&pg=1.

[82] ATSDR, Former Nansemond Ordnance Depot, Suffolk,Virginia, Public Health Assessments and Health Con-sultations, 2009, http://www.atsdr.cdc.gov/HAC/pha/pha.asp?docid=497&pg=3.

[83] T. F. Jenkins, J. C. Pennington, T. A. Ranney et al., “Char-acterization of explosives contamination at military firingranges,” ERDC TR-01-05, US Army Engineer Research andDevelopment Center, Hanover, NH, USA, 2001.

[84] US Environmental Protection Agency, “EPA SuperfundRecord of Decision: Bangor Naval Submarine Base, Sil-verdale, WA,” EPA/ROD/R10-94/107, 1994, http://www.epa.gov/superfund/sites/rods/fulltext/r1094107.pdf.

[85] ATSDR, Former Dupont Cladding Site at Cabin Lake (AkaDupont Barksdale Explosives Plant), WI, Public HealthAssessments and Health Consultations, 2009, http://www.atsdr.cdc.gov/HAC/pha/pha.asp?docid=750&pg=1.

[86] J. M. Brannon and T. E. Myers, “Review of fate and transportprocesses of explosives,” Tech. Rep. IRRP-97-2, U.S ArmyEngineer Waterways Experiment Station, Vicksburg, Miss,USA, 1997.

[87] US EPA, NPL Site Narrative for Pantex Plant (USDOE),PANTEX PLANT (USDOE) Pantex Village, Texas, 2011,http://www.epa.gov/superfund/sites/npl/nar1314.htm.

[88] US EPA, Joliet Army Ammunition Plant, Region 5 Super-fund, 2007, http://www.epa.gov/R5Super/fed fac/npl sites/ff npl jaap lap.html.

[89] B. Clark and R. Boopathy, “Evaluation of bioremediationmethods for the treatment of soil contaminated withexplosives in Louisiana Army Ammunition Plant, Minden,Louisiana,” Journal of Hazardous Materials, vol. 143, pp. 643–648, 2007.

[90] R. H. Gray and D. A. McGrath, “Environmental monitoringat DOE’s pantex plant in Amarillo, Texas,” Federal FacilitiesEnvironmental Journal, vol. 6, no. 1, pp. 79–88, 1995.

[91] M. E. Walsh, T. F. Jenkins, P. S. Schnitker, J. W. Elwell,and M. H. Stutz, “Evaluation of SW846 method 8330 forcharacterization of sites contaminated with residues of highexplosives,” CRREL Report 93-5, US Cold Regions Researchand Engineering Laboratory, Hanover, NH, USA, 1993.

[92] M. E. Walsh, A. D. Hewitt, M. R. Walsh et al., “Range char-acterization studies at Donnelly Training Area, Alaska: 2001and 2002,” Report ERDC/CRREL TR-04-3, US Army ColdRegions Research and Engineering Laboratory, Hanover, NH,USA, 2004.

[93] D. L. Pugh, “Milan Army Ammunition Plant contaminationsurvey,” USATHAMA Report DRXTH-FR-8213, US ArmyToxic and Hazardous Materials Agency, Aberdeen, MD, USA,1982.

[94] Pantex, http://www.pantex.com/.[95] B. D. Newman, D. D. Hickmott, and P. Gram, “Flow and

high explosives transport in a semiarid mesa in New Mexico,USA,” Vadose Zone Journal, vol. 6, pp. 774–785, 2007.

[96] Los Alamos National Laboratory, “RFI report for potentialrelease site 16-021(c),” LA-UR-98-4101, Los Alamos NationalLaboratory, Los Alamos, NM, USA, 1998.

[97] ATSDR, Savanna Army Depot Activity, Savanna, Illinois,Public Health Assessments and Health Consultations, 2009,http://www.atsdr.cdc.gov/HAC/pha/pha.asp?docid=601&pg=1#background.


[98] A. Eisentraeger, G. Reifferscheid, F. Dardenne, R. Blust, andA. Schofer, “Hazard characterization and identification of aformer ammunition site using microarrays, bioassays, andchemical analysis,” Environmental and Toxicological Chem-istry, vol. 26, pp. 634–646, 2007.

[99] US EPA, “Region 9 perchlorate update,” 1999, http://www.epa.gov/safewater/ccl/perchlor/r9699fac.pdf.

[100] E. T. Urbansky, M. L. Magnuson, C. A. Kelty, and S. K. Brown,“Perchlorate uptake by salt cedar (Tamarix ramosissima) inthe Las Vegas Wash riparian ecosystem,” Science of the TotalEnvironment, vol. 256, pp. 227–232, 2000.

[101] D. K. Tipton, D. E. Rolston, K. M. Scow et al., “Transportand biodegradation of perchlorate in soils,” Journal ofEnvironmental Quality, vol. 32, pp. 40–46, 2003.

[102] US AEC, “Unexploded ordnance (UXO),” 2011, http://aec.army.mil/usaec/technology/uxo00.html.

[103] Landmine and Cluster Munition Monitor, “Vietnam,”2003, http://www.the-monitor.org/index.php/publications/display?url=lm/2003/vietnam.html.

[104] BOMICO, “Tinh hinh o nhiem bom-min-vat no con sot laisau chien tranh,” (Situation on the Effects of Landmines,Bombs and Explosives Remaining after the War), draftpaper provided by VVAF, p. 7, as cited in: Landmine andCluster Munition Monitor, “Vietnam,” 2003, http://www.themonitor.org/index.php/publications/display?url=lm/2003/vietnam.html.

[105] D. Holdridge, “Viet Nam, US organisations to co-operate onmine survey,” Vietnam News Service, 29 January 2003.

[106] Interview with Nguyen Quang Vinh, Director, and Amb.Nguyen Quy Binh, Vice-Director of the BoundariesCommittee, Ministry of Foreign Affairs, Hanoi, as cited in:Landmine and Cluster Munition Monitor, “Vietnam,” 2003,http://www.the-monitor.org/index.php/publications/display?url=lm/2003/vietnam.html.

[107] T. F. Jenkins, M. E. Walsh, P. H. Miyares et al., Analysisof Explosives-Related Chemical Signatures in Soil SamplesCollected Near Buried Land Mines, Engineer Research andDevelopment Center, Hanover, NH, USA, 2000.

[108] Lao National Unexploded Ordnance Programme, “TheUnexploded Ordnance (UXO) Problem,” 2008, http://www.uxolao.org/uxo%20problem.html.

[109] C. Bendinelli, “Effects of land mines and unexploded ord-nance on the pediatric population and comparison withadults in rural Cambodia,” World Journal of Surgery, vol. 33,no. 5, pp. 1070–1074, 2009.

[110] Landmine and Cluster Munition Monitor, Cambodia, 2009http://www.the-monitor.org/index.php/publications/display?act=submit&pqs year=2002&pqs type=lm&pqs report=cambodia&pqs section=%23Heading3334#Heading3334.

[111] Australian Department of Defense, “UXO contaminationsites—by state,” 2010, http://www.defence.gov.au/uxo/drnreports/uxo by state.asp.

[112] UNICEF, “Landmines and UXO in Somaliland, Puntlandand Central & Southern Somalia: a feasibility study,” 2000,http://www.unicef.org/evaldatabase/files/SOM 00-003.pdf.

[113] H. Hafidh, “U.N. wants Iraq to issue more visas for its staff,”Tech. Rep., Reuters, Baghdad, Iraq, 2002.

[114] New York Times, “Israeli bomblets plague Lebanon,”2006, http://www.nytimes.com/2006/10/06/world/middleeast/06cluster.html?pagewanted=1& r=1&ref=todayspaper.

[115] A. Shadid, “In Lebanon, a war’s lethal harvest,” WashingtonPost Foreign Service, September 26, 2006, http://www.washingtonpost.com/wp-dyn/content/article/2006/09/25/AR2006092501500.html.

[116] Landmine and Cluster Munition Monitor, “Palestine,” 2009,http://www.the-monitor.org/index.php/publications/display?url=lm/2003/palestine.html.

[117] N. Prevost and UNICEF, “Unexploded Ordnance and MineAction in the Occupied Palestinian Territory,” Landmine andCluster Munition Monitor, 2003, http://www.the-monitor.org/index.php/publications/display?url=lm/2003/palestine.html.

[118] Landmine and Cluster Munition Monitor, “Armenia,” 2002,http://www.the-monitor.org/index.php/publications/display?act=submit&pqs year=2002&pqs type=lm&pqs report=ar-menia&pqs section=%23Heading14227#Heading14227.

[119] Landmine and Cluster Munition Monitor, “Chechnya,” 2009,http://www.the-monitor.org/index.php/publications/display?url=lm/2003/chechnya.html.

[120] Interfax, Unexploded Federal Ammunition Makes Up Mostof Landmines Used by Chechen Guerillas, Interfax, Moscow,Russia, 2003.

[121] “Unexploded federal ammunition makes up most of land-mines used by Chechen guerrillas,” Izvestia Interfax-AVN(Moscow), 20 May, 2003, Interview with Major YevgenyPasynok, Chief of Engineering Service, Grozny MilitaryCommandant’s Office.

[122] U. Khanbiev, Minister for Health of the Chechen Republic,citation translated from Russian by Landmine Monitor, 2002,http://www.chechenpress.com.

[123] O. Petrovsky, Hellish Surprises, UTRORU InformationAgency, 2002.

[124] J. A. MacDonald, M. J. Small, and M. G. Morgan, “Quan-tifying the risks of unexploded ordnance at closed militarybases,” Environmental Science and Technology, vol. 43, no. 2,pp. 259–265, 2009.

[125] Defense Science Board, Report of the Defense Science BoardTask Force on Unexploded Ordnance, Office of the Under Sec-retary of Defense for Acquisition, Technology, and Logistics,Washington, DC, USA, 2003.

[126] G. Ampleman, S. Thiboutot, S. Desilets, A. Gagnon, and A.Marois, “Evaluation of the soils contamination by explosivesat CFB Chilliwack and CFAD Rocky Point,” DREV ReportTR-2000-103, Quebec, Canada, 2000.

[127] G. Ampleman, D. Faucher, S. Thiboutot, J. Hawari, and F.Monteil-Rivera, “Evaluation of underwater contaminationby explosives and metals at Point Amour, Labrador and inthe Halifax Harbour Area,” DRDC Valcartier TR 2004-125,Defence Research and Development, Canada, 2004.

[128] T. F. Jenkins, A. D. Hewitt, C. L. Grant et al., “Identity anddistribution of residues of energetic compounds at army live-fire training ranges,” Chemosphere, vol. 63, no. 8, pp. 1280–1290, 2006.

[129] M. E. Walsh, C. M. Collins, R. N. Bailey, and C. L. Grant,“Composite sampling of sediments contaminated with whitephosphorus,” Special Report 97-30, US Army Cold RegionsResearch and Engineering Laboratory, Hanover, NH, USA,1997.

[130] T. F. Jenkins, C. L. Grant, M. E. Walsh et al., “Coping withspatial heterogeneity effects on sampling and analysis at anHMX-contaminated antitank firing range,” Field AnalyticalChemistry and Technology, vol. 3, no. 1, pp. 19–28, 1999.

[131] T. F. Jenkins, T. A. Ranney, A. D. Hewitt, M. E. Walsh, and K.L. Bjella, “Representative sampling for energetic compoundsat an anti-tank firing range,” ERDC/CRREL TR-04-7, USArmy Engineer Research and Development Center, Hanover,NH, USA, 2004.


[132] T. F. Jenkins, A. D. Hewitt, M. E. Walsh, C. L. Grant, andC. A. Ramsey, “Comment on data representativeness forrisk assessment by Rosemary Mattuck et al.,” Journal ofEnvironmental Forensics, vol. 6, no. 4, pp. 321–323, 2005.

[133] A. D. Hewitt, T. F. Jenkins, T. A. Ranney et al., “Estimates forexplosives residues from the detonation of army munitions,”ERDC/ CRREL TR-03-16, US Army Engineer Research andDevelopment Center, Hanover, NH, USA, 2003.

[134] AMEC, “Draft central impact area soil report technicalteam memorandum 01-13 Camp Edwards impact areagroundwater quality study, Camp Edwards, MassachusettsMilitary Reservation, Cape Cod, MA,” MMR-3915, AMECEarth and Environmental, Westford, Mass, USA, 2001.

[135] M.R. Walsh, S. Taylor, M. E. Walsh et al., “Residues from livefire detonations of 155-mm howitzer rounds,” ERDC/CRRELTR-05-14, US Army Engineer Research and DevelopmentCenter, Hanover, NH, USA, 2005.

[136] S. Taylor, A. Hewitt, J. Lever et al., “TNT particle sizedistributions from detonated 155-mm howitzer rounds,”Chemosphere, vol. 55, no. 3, pp. 357–367, 2004.

[137] J. Lewis, R. Martel, L. Trepanier, G. Ampleman, and S. Thi-boutot, “Quantifying the transport of energetic materials inunsaturated sediments from cracked unexploded ordnance,”Journal of Environmental Quality, vol. 38, no. 6, pp. 2229–2236, 2009.

[138] J. Clausen, J. Robb, D. Curry, and N. Korte, “A case studyof contaminants on military ranges: Camp Edwards, Mas-sachusetts, USA,” Environmental Pollution, vol. 129, no. 1, pp.13–21, 2004.

[139] J. L. Clausen, “Range assessment lessons learned,” FederalFacilities Environment Journal, vol. 16, pp. 49–62, 2005.

[140] W. F. Fitzpatrick, The Lessons of Massachusetts MilitaryReservation, Massachusetts Army National Guard, AEPI-IFP-1001B Army Environmental Policy Institute, Atlanta, Ga,USA, 2001.

[141] US EPA, “Administrative order for: Massachusetts MilitaryReservation training range and impact area,” EPA DocketRCRA 1-2001-0014, Washington, DC, USA, 2001.

[142] R. Boopathy, J. Manning, and C. F. Kulpa, “Optimizationof environmental factors for the biological treatment oftrinitrotoluene-contaminated soil,” Archives of Environmen-tal Contamination and Toxicology, vol. 32, no. 1, pp. 94–98,1997.

[143] USACHPPM, Draft site inspection No. 38-26-1339-95, SiteCS-19, Massachusetts Military Reservation, Cape Cod, Mas-sachusetts, US Center for Health Promotion and PreventiveMedicine, Aberdeen Proving Ground, Md, USA, 1994.

[144] T. F. Jenkins, M. E. Walsh, P. G. Thorne et al., “Assessmentof sampling error associated with collection and analysisof soil samples at a firing range contaminated with HMX,”CRREL Report 97-22, US Army Cold Regions Research andEngineering Laboratory, Hanover, NH, USA, 1997.

[145] S. Brochu, E. Diaz, S. Thiboutot et al., “Assessment of 100years of military training in Canada: the case of Cana-dian Force Base Petawawa,” TR-2008-118, Defense ResearchDevelopment Canada (DRDC-Valcartier), Quebec, Canada,2008.

[146] C. H. Racine, M. E. Walsh, C. M. Collins, D. J. Calkins,B. D. Roebuck, and L. Reitsma, “Waterfowl mortality inEagle River Flats, Alaska: the role of munition residues,”CRREL Report 92-5, US Army Cold Regions Research andEngineering Laboratory, Hanover, NH, USA, 1992.

[147] L. F. Phillips and C. A. Bouwkamp, “Effects of active firingrange activities on environmental media, Aberdeen Proving

Ground-Aberdeen Area, Maryland, 31 January—30 Decem-ber 1993,” Wastewater Management Study 32-24-HP16-94,US Army Environmental Hygiene Agency, Aberdeen ProvingGround, Md, USA, 1994.

[148] A. D. Hewitt and S. R. Bigl, “Elution of energetic compoundsfrom propellant and composition B residues,” ERDC/CRRELTR-05-13, Cold Regions Research and Engineering Labora-tory, US Army Engineer Research and Development Cen-ter, Hanover, NH, USA, 2005, http://libweb.erdc.usace.army.mil/Archimages/2705.pdf.

[149] G. Bordeleau, Etude hydrogeologique de la Base Aerienne deCold Lake, Alberta et determination de l’origine du nitratedans l’eau souterraine, M.S. thesis, INRSEte, Quebec, Canada,2007.

[150] D. M. Townsend and T. E. Myers, “Recent developments informulating model descriptors for subsurface transformationand sorption of TNT, RDX, and HMX,” Tech. Rep. IRRP-96-1, US Army Engineer Waterways Experiment Station,Vicksburg, Miss, USA, 1996.

[151] J. Wilkinson and D. Watt, Review of Demilitarization andDisposal Techniques for Munitions and Related Materials,Munitions Safety Information Analysis Centre (MSIAC),NATO Headquarters, Brussels, Belgium.

[152] T. A. Douglas, M. E. Walsh, C. J. McGrath, C. A. Weiss, A.M. Jaramillo, and T. P. Trainor, “Desorption of nitramineand nitroaromatic explosive residues from soils detonatedunder controlled conditions,” Environmental Toxicology andChemistry, vol. 30, no. 2, pp. 345–353, 2010.

[153] S. Taylor, J. H. Lever, J. Fadden, N. Perron, and B. Packer,“Simulated rainfall-driven dissolution of TNT, Tritonal,Comp B and Octol particles,” Chemosphere, vol. 75, no. 8, pp.1074–1081, 2009.

[154] K. S. Ro, A. Venugopal, D. D. Adrian et al., “Solubility of2,4,6-trinitrotoluene (TNT) in water,” Journal of Chemicaland Engineering Data, vol. 41, no. 4, pp. 758–761, 1996.

[155] J. M. Brannon, D. D. Adrian, J. C. Pennington, T. E. Myers,and C. A. Hayes, “Slow release of PCB, TNT, and RDX fromsoils and sediments,” Tech. Rep. EL-92-38, US Army Corps ofEngineers, Waterways Experiment Station, Vicksburg, Miss,USA, 1992.

[156] K. Verschueren, Handbook of Environmental Data on OrganicChemicals, D. Van Nostrand Reinhold Company, New York,NY, USA, 2nd edition, 1983.

[157] R. J. Spanggord, R. W. Mabey, T. W. Chou et al., “Environ-mental fate studies of HMX, phase II, detailed studies, finalreport,” SRI International, Menlo Park, Calif, USA, 1983.

[158] K. M. Dontsova, J. C. Pennington, C. Hayes, J. Simunek, andC. W. Williford, “Dissolution and transport of 2,4-DNT and2,6-DNT from M1 propellant in soil,” Chemosphere, vol. 77,no. 4, pp. 597–603, 2009.

[159] J. M. Phelan, J. V. Romero, J. L. Barnett, and D. R.Parker, “Solubility and dissolution kinetics of composition Bexplosive in water,” SAND Report SAND2002-2420, SandiaNational Laboratories, Albuquerque, NM, USA, 2002.

[160] J. C. Lynch, J. M. Brannon, and J. J. Delfino, “Effects ofcomponent interactions on the aqueous solubilities anddissolution rates of the explosive formulations octol, com-position B, and LX-14,” Journal of Chemical and EngineeringData, vol. 47, no. 3, pp. 542–549, 2002.

[161] J. C. Lynch, “Dissolution kinetics of high explosive com-pounds (TNT, RDX, HMX),” ERDC/EL TR-02-23, US ArmyEngineering Research and Development Center, Vicksburg,Miss, USA, 2002.


[162] J. C. Lynch, K. F. Myers, J. M. Brannon, and J. J. Delfino,“Effects of pH and temperature on the aqueous solubility anddissolution rate of 2,4,6-trinitrotoluene (TNT), hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX), and octahydro-1,3,5,7-tetranitro-1,3,5,7-tetrazocine (HMX),” Journal of Chemicaland Engineering Data, vol. 46, no. 6, pp. 1549–1555, 2001.

[163] J. H. Lever, S. Taylor, L. Perovich, K. Bjella, and B. Packer,“Dissolution of composition B detonation residuals,” Envi-ronmental Science and Technology, vol. 39, no. 22, pp. 8803–8811, 2005.

[164] J. C. Lynch, J. M. Brannon, and J. J. Delfino, “Dissolutionrates of three high explosive compounds: TNT, RDX, andHMX,” Chemosphere, vol. 47, no. 7, pp. 725–734, 2002.

[165] J. C. Lynch, J. M. Brannon, K. Hatfield, and J. J. Delfino,“An exploratory approach to modeling explosive compoundpersistence and flux using dissolution kinetics,” Journal ofContaminant Hydrology, vol. 66, no. 3-4, pp. 147–159, 2003.

[166] J. C. Pennington and J. M. Brannon, “Environmental fate ofexplosives,” Thermochimica Acta, vol. 384, no. 1-2, pp. 163–172, 2002.

[167] S. B. Haderlein, K. W. Weissmahr, and R. P. Schwarzen-bach, “Specific adsorption of nitroaromatic explosives andpesticides to clay minerals,” Environmental Science andTechnology, vol. 30, no. 2, pp. 612–622, 1996.

[168] K. A. Thorn, J. C. Pennington, and C. A. Hayes, “15NNMR investigation of the reduction and binding of TNTin an aerobic bench scale reactor simulating windrowcomposting,” Environmental Science and Technology, vol. 36,no. 17, pp. 3797–3805, 2002.

[169] H. Yamamoto, M. C. Morley, G. E. Speitel, and J. Clausen,“Fate and transport of high explosives in a sandy soil: adsorp-tion and desorption,” Soil and Sediment Contamination, vol.13, no. 5, pp. 459–477, 2004.

[170] J. L. Clausen, C. Scott, and I. Osgerby, “Fate of nitroglycerinand dinitrotoluene in soil at small arms training ranges,” Soiland Sediment Contamination, vol. 20, pp. 649–671, 2011.

[171] M. Windholz, The Merck Index, Merck and Company,Rahway, NJ, USA, 9th edition, 1979.

[172] J. Singh, S. D. Comfort, L. S. Hundal, and P. J. Shea, “Long-term RDX sorption and fate in soil,” Journal of EnvironmentalQuality, vol. 27, no. 3, pp. 572–577, 1998.

[173] J. C. Pennington and W. H. Patrick Jr., “Adsorption anddesorption of 2,4,6-trinitrotoluene by soils,” Journal ofEnvironmental Quality, vol. 19, no. 3, pp. 559–567, 1990.

[174] J. M. Brannon, C. B. Price, C. Hayes, and S. L. Yost, “Aquifersoil cation substitution and adsorption of TNT, RDX, andHMX,” Soil and Sediment Contamination, vol. 11, no. 3, pp.327–338, 2002.

[175] C. J. McGrath, “Review of formulations for processes affect-ing the subsurface transport of explosives,” Tech. Rep. IRRP-95-2, US Army Corps of Engineers, Waterways ExperimentStation, Vicksburg, Miss, USA, 1995.

[176] S. K. Xue, I. K. Iskandar, and H. M. Selim, “Adsorption-desorption of 2,4,6-trinitrotoluene and hexahydro-1,3,5-trinitro-1,3,5-triazine in soils,” Soil Science, vol. 160, no. 5,pp. 317–327, 1995.

[177] J. C. Pennington, D. Gunnison, D. W. Harrelson et al.,“Natural attenuation of explosives in soil and water systemsat Department of Defence sites: interim report,” Tech.Rep. EL-99-8, US Army Corps of Engineers, WaterwaysExperiment Station, Vicksburg, Miss, USA, 1999.

[178] T. E. Myers, J. M. Brannon, J. C. Pennington et al.,“Laboratory studies of soil sorption/transformation of TNT,RDX, and HMX,” Tech. Rep. IRRP-98-8, US Army EngineerWaterways Experiment Station, Vicksburg, Miss, USA, 1998.

[179] H. M. Selim and I. K. Iskandar, “Sorption-desorption andtransport of TNT and RDX in soils,” CRREL Report 94-7, USArmy Cold Regions Research and Engineering Laboratory,Hanover, NH, USA, 1994.

[180] C. C. Ainsworth, S. D. Harvey, J. E. Szecsodt et al., “Rela-tionship between the leachability characteristics of uniqueenergetic compounds and soil properties,” Project 91PP1800,US Army Biomedical Research and Development Laboratory,Fort Detrick, Md, USA, 1993.

[181] C. B. Price, J. M. Brannon, S. L. Yost et al., “Transformationof RDX and HMX under controlled Eh/pH conditions,”Tech. Rep. IRRP-98-2, US Army Engineering WaterwaysExperimental Station, Vicksburg, Miss, USA, 1998.

[182] J. Brannon, P. Deliman, C. Ruiz et al., “Conceptual modeland process descriptor formulations for fate and transportof UXO,” Tech. Rep. IRRP-99-1, US Army Engineer Researchand Development Center, Vicksburg, Miss, USA, 1999.

[183] T. W. Sheremata, S. Thiboutot, G. Ampleman, L. Paquet,A. Halasz, and J. Hawari, “Fate of 2,4,6-trinitrotolueneand its metabolites in natural and model soil systems,”Environmental Science and Technology, vol. 33, no. 22, pp.4002–4008, 1999.

[184] C. B. Price, J. M. Brannon, S. L. Yost, and A. H. Charolett,“Adsorption and transformation of explosives in low-carbonaquifer soils,” Report ERDC/EL TR-00-11, US Army Corpsof Engineers, Engineer Research and Development Center,Vicksburg, Miss, USA, 2000.

[185] M. V. Cattaneo, J. C. Pennington, J. M. Brannon, D. Gun-nison, D. W. Harrelson, and M. Zakikhani, “Natural attenu-ation of explosives,” in Remediation of Hazardous Waste Con-taminated Soils, Marcel Dekker, New York, NY, USA, 2000.

[186] N. Singh, D. Hennecke, J. Hoerner, W. Koerdel, and A.Schaeffer, “Sorption-desorption of trinitrotoluene in soils:effect of saturating metal cations,” Bulletin of EnvironmentalContamination and Toxicology, vol. 80, no. 5, pp. 443–446,2008.

[187] K. W. Weissmahr, S. B. Haderlein, and R. P. Schwarzenbach,“Complex formation of soil minerals with nitroaromaticexplosives and other π-acceptors,” Soil Science Society ofAmerica Journal, vol. 62, no. 2, pp. 369–378, 1998.

[188] J. M. Brannon, C. B. Price, S. L. Yost, C. Hayes, J. E. Mirecki,and B. Porter, “Fate and transport parameters for firing rangeresidues,” In Distribution and Fate of Energetics on DoDTest and Training Ranges: Interim Report 4 ERDC TR-04-4, US Army Engineer Research and Development Center,Vicksburg, Miss, USA, 2004.

[189] W. R. Haag, R. Spanggord, T. Mill et al., “Aquatic environ-mental fate of nitroguanidine,” Environmental Toxicology andChemistry, vol. 9, no. 11, pp. 1359–1367, 1990.

[190] J. L. Clausen, C. Scott, N. Mulherin et al., “Sorption/desorp-tion measurements of nitroglycerin and dinitrotoluene inCamp Edwards, Massachusetts soil,” ERDC/CRREL TR-10-1, Cold Regions Research and Engineering Laboratory, USArmy Engineer Research and Development Center, Hanover,NH, USA, 2010.

[191] G. Ji and X. Kong, “Adsorption of chloride, nitrate, andperchlorate by variable charge soils,” Pedosphere, vol. 2, pp.317–326, 1992.


[192] F. Monteil-Rivera, C. Groom, and J. Hawari, “Sorptionand degradation of octahydro-1,3,5,7-tetranitro-1,3,5,7-tetrazocine in soil,” Environmental Science and Technology,vol. 37, no. 17, pp. 3878–3884, 2003.

[193] C. B. Price, J. M. Brannon, and C. A. Hayes, “Effect ofredox potential and pH on TNT transformation in soil-waterslurries,” Journal of Environmental Engineering, vol. 123, no.10, pp. 988–992, 1997.

[194] T. A. Douglas, M. E. Walsh, C. J. McGrath, and C. A. Weiss,“Investigating the fate of nitroaromatic (TNT) and nitramine(RDX and HMX) explosives in fractured and pristine soils,”Journal of Environmental Quality, vol. 38, no. 6, pp. 2285–2294, 2009.

[195] D. J. Glover and J. C. Hoffsommer, “Photolysis of RDX.Identification and reactions of products,” Tech. Rep. NSWCTR-79-349, Naval Surface Weapons Center, Silver Spring,Md, USA, 1979.

[196] N. E. Burlinson, L. A. Kaplan, and C. E. Adams, “Photochem-istry of TNT: investigation of the “pink water” problem,”NOLTR 73-172, Naval Ordnance Laboratory, Silver Spring,Md, USA, 1973.

[197] M. Godejohann, A. Preiss, K. Levsen, K.-M. Wollin, and C.Mugge, “Determination of polar organic pollutants in aque-ous samples of former ammunition sites in Lower Saxonyby means of HPLC/photodiode array detection (HPLC/PDA)and proton nuclear magnetic resonance spectroscopy (1H-NMR),” Acta Hydrochimica et Hydrobiologica, vol. 26, no. 6,pp. 330–337, 1998.

[198] T. C. Schmidt, M. Petersmann, L. Kaminski, E. V. Low, andG. Stork, “Analysis of aminobenzoic acids in waste waterfrom a former ammunition plant with HPLC and combineddiode array and fluorescence detection,” Fresenius’ Journal ofAnalytical Chemistry, vol. 357, no. 1, pp. 121–126, 1997.

[199] A. Preiss, M. Elend, S. Gerling, E. Berger-Preiss, and K.Steinbach, “Identification of highly polar nitroaromatic com-pounds in leachate and ground water samples from a TNT-contaminated waste site by LC-MS, LC-NMR, and off-lineNMR and MS investigations,” Analytical and BioanalyticalChemistry, vol. 389, no. 6, pp. 1979–1988, 2007.

[200] D. Hennecke, W. Kordel, K. Steinbach, and B. Her-rmann, “Transformation processes of explosives in naturalwater/sediment systems,” in Proceedings of the 10th Interna-tional UFZ Deltares/TNO Conference on Management of Soil,Groundwater and Sediments, Milano, Italy, September 2008.

[201] D. T. Burton and S. D. Turley, “Reduction of hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX) toxicity to the cladoceranCeriodaphnia dubia following photolysis in sunlight,” Bul-letin of Environmental Contamination and Toxicology, vol. 55,no. 1, pp. 89–95, 1995.

[202] P. Bose, W. H. Glaze, and D. S. Maddox, “Degradation ofRDX by various advanced oxidation processes: II. organicby-products,” Water Research, vol. 32, no. 4, pp. 1005–1018,1998.

[203] G. R. Peyton, M. H. LeFaivre, and S. W. Maloney, “Veri-fication of RDX photolysis mechanism,” Report CERL-TR-99/93, Construction Engineering Research Lab (Army), IL,1999.

[204] H. M. Seth-Smith, Microbial degradation of RDX, PhDDissertation, University of Cambridge, UK, 2002.

[205] M. S. Simmons and R. G. Zepp, “Influence of humicsubstances on photolysis of nitroaromatic compounds inaqueous systems,” Water Research, vol. 20, no. 7, pp. 899–904,1986.

[206] A. Mills, A. Seth, and G. Peters, “Alkaline hydrolysis oftrinitrotoluene, TNT,” Physical Chemistry Chemical Physics,vol. 5, no. 18, pp. 3921–3927, 2003.

[207] J. L. Davis, M. C. Brooks, S. L. Larson, C. C. Nestler, andD. R. Felt, “Lime treatment of explosives-contaminated soilfrom munitions plants and firing ranges,” Soil and SedimentContamination, vol. 15, no. 6, pp. 565–580, 2006.

[208] M. Emmrich, “Kinetics of the alkaline hydrolysis of 2,4,6-trinitrotoluene in aqueous solution and highly contaminatedsoils,” Environmental Science and Technology, vol. 33, no. 21,pp. 3802–3805, 1999.

[209] A. Saupe, H. J. Garvens, and L. Heinze, “Alkaline hydrolysisof TNT and TNT in soil followed by thermal treatment ofthe hydrolysates,” Chemosphere, vol. 36, no. 8, pp. 1725–1744,1998.

[210] R. Bajpai, D. Parekh, S. Herrmann, M. Popovic, J. Paca,and M. Qasim, “A kinetic model of aqueous-phase alkalihydrolysis of 2,4,6-trinitrotoluene,” Journal of HazardousMaterials, vol. 106, no. 1, pp. 37–44, 2004.

[211] R. J. Spanggord, T. Mill, T. W. Chou, R. W. Mabey, J. H.Smith, and S. Lee, “Environmental fate studies on certainmunition wastewater constituents,” Final Report, Phase 1:Literature Review LSU-7934, SRI International, Menlo Park,Calif, USA, 1980.

[212] R. J. Spanggord, R. W. Mabey, T. W. Chou et al., “Envi-ronmental fate studies of HMX, screening studies, finalreport, phase I—Laboratory study,” SRI Project LSU-4412,SRI International, Menlo Park, CA, for US Army MedicalResearch and Development Command, Fort Detrick, Md,USA, 1982.

[213] V. K. Balakrishnan, A. Halasz, and J. Hawari, “Alkalinehydrolysis of the cyclic nitramine explosives RDX, HMX,and CL-20: new insights into degradation pathways obtainedby the observation of novel intermediates,” EnvironmentalScience and Technology, vol. 37, pp. 1838–1843, 2003.

[214] K. B. Gregory, P. Larese-Casanova, G. F. Parkin, and M. M.Scherer, “Abiotic transformation of hexahydro-1,3,5-trinito-1,3,5-triazine by Fe II bound to magnetite,” EnvironmentalScience and Technology, vol. 38, no. 5, pp. 1408–1414, 2004.

[215] D. M. Townsend, T. E. Myers, and D. D. Adrian, “2,4,6-trinitrotoluene (TNT) transformation/sorption in thin-disksoil columns,” Tech. Rep. IRRP-95-4, US Army EngineerWaterways Experiment Station, Vicksburg, Miss, USA, 1995.

[216] H. M. Selim, S. K. Xue, and I. K. Iskandar, “Transportof 2,4,6-trinitrotoluene and hexahydro-1,3,5-trinitro-1,3,5-triazine in soils,” Soil Science, vol. 160, no. 5, pp. 328–339,1995.

[217] D. L. Kaplan and A. M. Kaplan, “Thermophilic biotransfor-mations of 2,4,6-trinitrotoluene under simulated compost-ing conditions,” Applied and Environmental Microbiology, vol.44, no. 3, pp. 757–760, 1982.

[218] N. G. McCormick, F. E. Feeherry, and H. S. Levin-son, “Microbial transformation of 2,4,6-trinitrotoluene andother nitroaromatic compounds,” Applied and EnvironmentalMicrobiology, vol. 31, no. 6, pp. 949–958, 1976.

[219] C. B. Price, J. M. Brannon, and C. A. Hayes, “Transformationof 2,4,6-trinitrotoluene under controlled Eh/pH conditions,”Tech. Rep. IRRP-95-5, US Army Engineer Waterways Exper-iment Station, Vicksburg, Miss, USA, 1995.

[220] R. G. Reifler and B. F. Smets, “Enzymatic reduction of 2,4,6-trinitrotoluene and related nitroarenes: kinetics linked toone-electron redox potentials,” Environmental Science andTechnology, vol. 34, no. 18, pp. 3900–3906, 2000.


[221] J. Z. Bandstra, R. Miehr, R. L. Johnson, and P. G. Tratnyek,“Reduction of 2,4,6-trinitrotoluene by iron metal: kineticcontrols on product distributions in batch experiments,”Environmental Science and Technology, vol. 39, no. 1, pp. 230–238, 2005.

[222] J. F. Devlin, J. Klausen, and R. P. Schwarzenbach, “Kineticsof nitroaromatic reduction on granular iron in recirculatingbatch experiments,” Environmental Science and Technology,vol. 32, no. 13, pp. 1941–1947, 1998.

[223] J. M. Brannon, C. B. Price, and C. Hayes, “Abiotic trans-formation of TNT in montmorillonite and soil suspensionsunder reducing conditions,” Chemosphere, vol. 36, no. 6, pp.1453–1462, 1998.

[224] R. L. Crawford, “Biodegradation of nitrated munitions com-pounds and herbicides by obligately anaerobic bacteria,” inBiodegradation of Nitroaromatic Compounds, J. C. Spain, Ed.,Plenum Press, New York, NY, USA, 1995.

[225] S. B. Funk, D. J. Roberts, D. L. Crawford, and R. L. Crawford,“Initial-phase optimization for bioremediation of munitioncompound—contaminated soils,” Applied and Environmen-tal Microbiology, vol. 59, no. 7, pp. 2171–2177, 1993.

[226] C. B. Price, J. M. Brannon, S. L. Yost, and C. A. Hayes,“Relationship between redox potential and pH on RDXtransformation in soil-water slurries,” Journal of Environmen-tal Engineering, vol. 127, no. 1, pp. 26–31, 2001.

[227] N. G. McCormick, J. H. Cornell, and A. M. Kaplan,“Biodegradation of hexahydro-1,3,5-trinitro-1,3,5-triazine,”Applied and Environmental Microbiology, vol. 42, no. 5, pp.817–823, 1981.

[228] D. L. Kaplan, “Biotechnology and bioremediation for organicenergetic compounds,” in Organic Energetic Compounds, P.Marikas, Ed., pp. 373–395, Nova Science Publishers, NewYork, NY, USA, 1996.

[229] J. Park, S. D. Comfort, P. J. Shea, and T. A. Machacek, “Reme-diating munitions-contaminated soil with zero valent ironand cationic surfactants,” Journal of Environmental Quality,vol. 33, no. 4, pp. 1305–1313, 2004.

[230] L. S. Hundal, J. Singh, E. L. Bier, P. J. Shea, S. D. Comfort,and W. L. Powers, “Removal of TNT and RDX from waterand soil using iron metal,” Environmental Pollution, vol. 97,no. 1-2, pp. 55–64, 1997.

[231] D. Colon, E. J. Weber, J. L. Anderson, P. Winget, and L. A.Suarez, “Reduction of nitrosobenzenes and n-hydroxyl-anilines by Fe(II) species: elucidation of the reaction mech-anism,” Environmental Science and Technology, vol. 40, pp.4449–4454, 2006.

[232] D. Liu, K. Thomson, and A. C. Anderson, “Identification ofnitroso compounds from biotransformation of 2,4-dinitro-toluene,” Applied and Environmental Microbiology, vol. 47,no. 6, pp. 1295–1298, 1984.

[233] J. S. Zhao, D. Fournier, S. Thiboutot, G. Ampleman, and J.Hawari, “Biodegradation and bioremediation of explosives,”in Applied Bioremediation and Phytoremediation, A. Singhand O. P. Ward, Eds., Springer, New York, NY, USA, 2004.

[234] M. B. Pasti-Grigsby, A. Paszczynski, S. Goszczynski, D. L.Crawford, and R. L. Crawford, “Influence of aromatic sub-stitution patterns on azo dye degradability by Streptomycesspp. and Phanerochaete chrysosporium,” Applied and Environ-mental Microbiology, vol. 58, no. 11, pp. 3605–3613, 1992.

[235] J. T. Spadaro, M. H. Gold, and V. Renganathan, “Degradationof azo dyes by the lignin-degrading fungus Phanerochaetechrysosporium,” Applied and Environmental Microbiology, vol.58, no. 8, pp. 2397–2401, 1992.

[236] G. Soli, “Microbial degradation of cyclonite (RDX),” ReportNWC-TP-5525 / AD-762 751, US National Technical Infor-mation Service (Naval Weapons Center China Lake CA),Washington, DC, USA, 1973.

[237] R. J. Spanggord, T. Mill, T. W. Chou, R. W. Mabey, W. H.Smith, and S. Lee, “Environmental fate studies on certainmunition wastewater constituents, final report, part II—laboratory study,” AD A099256, SRI International, MenloPark, CA, for US Army Medical Research and DevelopmentCommand, Fort Detrick, Md, USA, 1980.

[238] N. V. Coleman, D. R. Nelson, and T. Duxbury, “Aero-bic biodegradation of hexahydro-1,3,5 trinitro-1,3,5-triazine(RDX) as a nitrogen source by a Rhodococcus sp., strainDN22,” Soil Biology and Biochemistry, vol. 30, no. 8-9, pp.1159–1167, 1998.

[239] N. G. McCormick, J. H. Cornell, and A. M. Kaplan, “Theanaerobic transformation of RDX and HMX and theiracetylated derivatives,” Tech. Rep. A149464 (TR85–008), USArmy Natick Research and Development Center, Natick,Mass, USA, 1985.

[240] K. M. Regan and R. L. Crawford, “Characterization ofClostridium bifermentans and its biotransformation of2,4,6-trinitrotoluene (TNT) and 1,3,5-triaza-1,3,5-trinitro-cyclohexane (RDX),” Biotechnology Letters, vol. 16, no. 10, pp.1081–1086, 1994.

[241] C. A. Groom, S. Beaudet, A. Halasz, L. Paquet, and J. Hawari,“Detection of the cyclic nitramine explosives hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX) and octahydro-1,3,5,7-tetranitro-1,3,5,7-tetrazine (HMX) and their degradationproducts in soil environments,” Journal of ChromatographyA, vol. 909, no. 1, pp. 53–60, 2001.

[242] C. F. Shen, J. A. Hawari, L. Paquet, G. Ampleman, S.Thiboutot, and S. R. Guiot, “Explosive biodegradation in soilslurry batch reactors amended with exogenous microorgan-isms,” Water Science and Technology, vol. 43, no. 3, pp. 291–298, 2001.

[243] C. F. Shen, S. R. Guiot, S. Thiboutot, G. Ampleman, and J.Hawari, “Fate of explosives and their metabolites in bioslurrytreatment processes,” Biodegradation, vol. 8, no. 5, pp. 339–347, 1997.

[244] S. Toze and L. Zappia, “Microbial degradation of munitioncompounds in production wastewater,” Water Research, vol.33, no. 13, pp. 3040–3045, 1999.

[245] D. L. Freedman and K. W. Sutherland, “Biodegrada-tion of hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX) undernitrate-reducing conditions,” Water Science and Technology,vol. 38, no. 7, pp. 33–40, 1998.

[246] R. Boopathy, M. Gurgas, J. Ullian, and J. F. Manning,“Metabolism of explosive compounds by sulfate-reducingbacteria,” Current Microbiology, vol. 37, no. 2, pp. 127–131,1998.

[247] R. Boopathy, J. Manning, and C. F. Kulpa, “Biotransfor-mation of explosives by anaerobic consortia in liquid cul-ture and in soil slurry,” International Biodeterioration andBiodegradation, vol. 41, no. 1, pp. 67–74, 1998.

[248] J. Hawari, A. Halasz, T. Sheremata et al., “Characteriza-tion of metabolites during biodegradation of hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX) with municipal anaerobicsludge,” Applied and Environmental Microbiology, vol. 66, no.6, pp. 2652–2657, 2000.

[249] C. L. Kitts, D. P. Cunningham, and P. J. Unkefer, “Isolationof three hexahydro-1,3,5-trinitro-1,3,5-triazine-degradingspecies of the family Enterobacteriaceae from nitramine


explosive contaminated soil,” Applied and EnvironmentalMicrobiology, vol. 60, no. 12, pp. 4608–4711, 1994.

[250] D. M. Young, P. J. Unkefer, and K. L. Ogden, “Biotransforma-tion of hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX) by aprospective consortium and its most effective isolate Serratiamarcescens,” Biotechnology and Bioengineering, vol. 53, pp.515–522, 1997.

[251] T. W. Sheremata and J. Hawari, “Mineralization of RDXby the white rot fungus Phanerochaete chrysosporium tocarbon dioxide and nitrous oxide,” Environmental Scienceand Technology, vol. 34, no. 16, pp. 3384–3388, 2000.

[252] P. R. Binks, S. Nicklin, and N. C. Bruce, “Degradationof hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX) by Steno-trophomonas maltophilia PB1,” Applied and EnvironmentalMicrobiology, vol. 61, no. 4, pp. 1318–1322, 1995.

[253] B. Bhushan, A. Halasz, S. Thiboutot, G. Ampleman, andJ. Hawari, “Chemotaxis-mediated biodegradation of cyclicnitramine explosives RDX, HMX, and CL-20 by Clostridiumsp. EDB2,” Biochemical and Biophysical Research Communi-cations, vol. 316, no. 3, pp. 816–821, 2004.

[254] Y. Yang, X. Wang, P. Yin, W. Li, and P. Zhou, “Studies on threestrains of Corynebacterium degrading cyclotrimethylene-trinitroamine (RDX),” Acta Microbiologica Sinica, vol. 23, pp.251–256, 1983.

[255] S. Thiboutot, J. Lavigne, G. Ampleman et al., The Inter-national Symposium on Energetic Materials Technology,Orlando, Fla, USA, 1994.

[256] A. M. Jones, S. Labelle et al., “Assessment of the aerobicbiodegradation potential of RDX, TNT, GAP, and NC,”in Environmental Biotechnology—Principles and Applications,M. Moo-Young, W. A. Anderson, and A. M. Chakrabarty,Eds., pp. 368–381, Kluwer Academic Press, New York, NY,USA, 1995.

[257] B. van Aken, J. M. Yoon, C. L. Just, and J. L. Schnoor, “Meta-bolism and mineralization of hexahydro-1,3,5-trinitro-1,3,5-triazine inside poplar tissues (Populus deltodesp × nigra DN-34),” Environmental Science and Technology, vol. 38, no. 17,pp. 4572–4579, 2004.

[258] C. Axtell, C. G. Johnston, and J. A. Bumpus, “Bioremediationof soil contaminated with explosives at the Naval WeaponsStation Yorktown,” Soil and Sediment Contamination, vol. 9,no. 6, pp. 537–548, 2000.

[259] D. Fournier, A. Halasz, S. Thiboutot, G. Ampleman, D.Manno, and J. Hawari, “Biodegradation of octahydro-1,3,5,7-tetranitro-1,3,5,7-tetrazocine (HMX) by Phanero-chaete chrysosporium: new insight into the degradation path-way,” Environmental Science and Technology, vol. 38, no. 15,pp. 4130–4133, 2004.

[260] J. S. Zhao, J. Spain, and J. Hawari, “Phylogenetic andmetabolic diversity of hexahydro-1,3,5-trinitro-1,3,5- tri-azine (RDX)-transforming bacteria in strictly anaerobicmixed cultures enriched on RDX as nitrogen source,” FEMSMicrobiology Ecology, vol. 46, no. 2, pp. 189–196, 2003.

[261] R. Boopathy, “Enhanced biodegradation of cyclotetramethyl-enetetranitramine (HMX) under mixed electron-acceptorcondition,” Bioresource Technology, vol. 76, no. 3, pp. 241–244, 2001.

[262] A. Haıdour and J. L. Ramos, “Identification of productsresulting from the biological reduction of 2,4,6-trinitro-toluene, 2,4-dinitrotoluene, and 2,6-dinitrotoluene by Pseu-domonas sp.,” Environmental Science and Technology, vol. 30,no. 7, pp. 2365–2370, 1996.

[263] U. Lendenmann, J. C. Spain, and B. F. Smets, “Simultaneousbiodegradation of 2,4-dinitrotoluene and 2,6-dinitrotoluenein an aerobic fluidized-bed biofilm reactor,” EnvironmentalScience and Technology, vol. 32, no. 1, pp. 82–87, 1998.

[264] S. F. Nishino, G. Paoli, and J. C. Spain, “Aerobic biodegrada-tion of dinitrotoluenes and pathway for bacterial degradationof 2,6-dinitrotoluene,” Applied and Environmental Microbiol-ogy, vol. 66, pp. 2138–2147, 2000.

[265] N. G. McCormick, J. H. Cornell, and A. M. Kaplan, “Iden-tification of biotransformation products from 2,4 dinitro-toluene,” Applied and Environmental Microbiology, vol. 35,no. 5, pp. 945–948, 1978.

[266] K. F. Reardon and J. C. Spain, “Immobilized cell bioreactorfor 2,4-dinitrotoluene degradation,” in Proceedings of theABSTR Biot86, 205th ACS National Meeting, AmericanChemical Society, Washington, DC, USA, 1993.

[267] R. E. Williams and N. C. Bruce, “The role of nitrate esterreductase enzymes in the biodegradation of explosives,” inBiodegradation of Nitroaromatic Compounds and Explosives,J. C. Spain, J. B. Hughes, and H.-J. Knackmuss, Eds., LewisPublishers, Boca Raton, Fla, USA, 2000.

[268] G. F. White, J. R. Snape, and S. Nicklin, “Biodegradationof glycerol trinitrate and pentaerythritol tetranitrate by Agro-bacterium radiobacter,” Applied and Environmental Microbiol-ogy, vol. 62, no. 2, pp. 637–642, 1996.

[269] M. Meng, W. Q. Sun, L. A. Goelhaar et al., “Denitration ofglycerol trinitrate by resting cells and cell extracts of Bacillusthuringiensis/cereus and Enterobacter agglemerans,” Appliedand Environmental Microbiology, vol. 61, no. 7, pp. 2548–2553, 1995.

[270] T. M. Wendt, J. H. Cornell, and A. M. Kaplan, “Microbialdegradation of glycerol nitrates,” Applied and EnvironmentalMicrobiology, vol. 36, no. 5, pp. 693–699, 1978.

[271] S. Bhaumik, C. Christodoulatos, G. P. Korfiatis, and B.W. Brodman, “Aerobic and anaerobic biodegradation ofnitroglycerin in batch and packed bed bioreactors,” WaterScience and Technology, vol. 36, no. 2-3, pp. 139–146, 1997.

[272] C. Christodoulatos, S. Bhaumik, and B. W. Brodman,“Anaerobic biodegradation of nitroglycerin,” Water Research,vol. 31, no. 6, pp. 1462–1470, 1997.

[273] C. E. French, S. Nicklin, and N. C. Bruce, “Sequence andproperties of pentaerythritol tetranitrate reductase fromEnterobacter cloacae PB2,” Journal of Bacteriology, vol. 178,no. 22, pp. 6623–6627, 1996.

[274] J. V. Accashian, R. T. Vinopal, B.-J. Kim, and B. F. Smets, “Aer-obic growth on nitroglycerin as the sole carbon, nitrogen,and energy source by a mixed bacterial culture,” Applied andEnvironmental Microbiology, vol. 64, no. 9, pp. 3300–3304,1998.

[275] G. F. White, J. R. Snape, and S. Nicklin, “Bacterial biodegra-dation of glycerol trinitrate,” International Biodeteriorationand Biodegradation, vol. 38, no. 2, pp. 77–82, 1996.

[276] D. S. Blehert, B. G. Fox, and G. H. Chambliss, “Cloningand sequence analysis of two Pseudomonas flavoproteinxenobiotic reductases,” Journal of Bacteriology, vol. 181, no.20, pp. 6254–6263, 1999.

[277] J. R. Snape, N. A. Walkley, A. P. Morby, S. Nicklin, and G.F. White, “Purification, properties, and sequence of glyceroltrinitrate reductase from Agrobacterium radiobacter,” Journalof Bacteriology, vol. 179, no. 24, pp. 7796–7802, 1997.

[278] D. S. Blehert, K. Becker, and G. H. Chambliss, “Isolationand characterization of bacteria that degrade nitroglycerin,”in Proceedings of the Tri-Service Environmental Technology


Workshop. Enhancing Readiness through Environmental Qual-ity Technology, pp. 197–204, Defense Technical InformationCenter, Hershey, Pa, USA, May 1996, ADP017715.

[279] S. J. Marshall and G. F. White, “Complete denitration ofnitroglycerin by bacteria isolated from a washwater soak-away,” Applied and Environmental Microbiology, vol. 67, no.6, pp. 2622–2626, 2001.

[280] S. Yost, Effects of redox potential and pH on the fate of nitro-glycerin in a surface and aquifer soil, M.S. thesis, LouisianaState University, Baton Rouge, La, USA, 2004.

[281] B. van Aken and J. L. Schnoor, “Evidence of perchlorate(ClO−

4 ) reduction in plant tissues (Poplar tree) using radio-labeled 36ClO4

−,” Environmental Science and Technology, vol.36, no. 12, pp. 2783–2788, 2002.

[282] G. B. Rikken, A. G. M. Kroon, and C. G. Ginkel, “Trans-formation of (per)chlorate into chloride by a newly isolatedbacterium: reduction and dismutation,” Applied Microbiologyand Biotechnology, vol. 45, no. 3, pp. 420–426, 1996.

[283] J. D. Coates, U. Michaelidou, R. A. Bruce, S. M. O’Connor, J.N. Crespi, and L. A. Achenbach, “Ubiquity and diversity ofdissimilatory (per)chlorate-reducing bacteria,” Applied andEnvironmental Microbiology, vol. 65, no. 12, pp. 5234–5241,1999.

[284] V. N. Korenkov, V. I. Romanenko, S. I. Kuznetsov, and J. V.Voronnov, “Process for purification of industrial waste watersfrom perchlorates and chlorates,” US patent 3,943,055, 1976.

[285] W. Wallace, T. Ward, A. Breen, and H. Attaway, “Identifica-tion of an anaerobic bacterium which reduces perchlorateand chlorate as Wolinella succinogenes,” Journal of IndustrialMicrobiology, vol. 16, no. 1, pp. 68–72, 1996.

[286] R. A. Bruce, L. A. Achenbach, and J. D. Coates, “Reduction of(per)chlorate by a novel organism isolated from paper millwaste,” Environmental Microbiology, vol. 1, no. 4, pp. 319–329, 1999.

[287] D. C. Herman and W. T. Frankenberger, “Microbial-mediated reduction of perchlorate in groundwater,” Journalof Environmental Quality, vol. 27, no. 4, pp. 750–754, 1998.

[288] B. E. Logan, H. Zhang, P. Mulvaney, M. G. Milner, I. M. Head,and R. F. Unz, “Kinetics of perchlorate—and chlorate—respiring bacteria,” Applied and Environmental Microbiology,vol. 67, no. 6, pp. 2499–2506, 2001.

[289] B. C. Okeke, T. Giblin, and W. T. Frankenberger, “Reductionof perchlorate and nitrate by salt tolerant bacteria,” Environ-mental Pollution, vol. 118, no. 3, pp. 357–363, 2002.


Research Article

Occurrence of Vanadium in Belgian and European Alluvial Soils

V. Cappuyns and E. Slabbinck

Center for Economics and Corporate Sustainability, Hogeschool-Universiteit Brussel, Stormstraat 2, 1000 Brussels, Belgium

Correspondence should be addressed to V. Cappuyns, [email protected]

Received 16 November 2011; Accepted 23 January 2012


Copyright © 2012 V. Cappuyns and E. Slabbinck. This is an open access article distributed under the Creative CommonsAttribution License, which permits unrestricted use, distribution, and reproduction in any medium, provided the original work isproperly cited.

Vanadium (V) is a naturally occurring trace element, but total concentrations in soils and sediments are also dependent on theparent material and might be influenced by anthropogenic activities (e.g., steel industry). Despite the fact that threshold values forV in soils and/or sediments exist in various European countries, in Belgium, V is not taken into account when the environmentalquality of soils and sediments has to be evaluated, despite the existence of several (diffuse) sources for V. In the first part of thestudy, the occurrence of V alluvial soils in Belgium was compared with V concentrations in alluvial soils (floodplain soils) acrossEurope. By analysis of both the Belgian and European data, the relationship between physicochemical soil characteristics and totalV concentrations was quantified and some areas polluted with V were detected. A regression equation, in which V concentrationsin alluvial soils were expressed as a function of major element composition, was proposed for the Belgian and European data.Additionally, single extractions with CaCl2 (0.01 mol L−1) and ammonium-EDTA (0.05 mol L−1) were used to estimate short- andlong-term mobility of V in the alluvial soils.

1. Introduction

1.1. Vanadium in Soils and Sediments. Vanadium (V) natu-rally occurs as trace element in soils and sediments. The aver-age concentration in soil is estimated to be around 150 μg g−1

[1] and 90 μg g−1 [2]. The concentration of V in soils andsediments also depends on the parent material and the occur-rence of V-containing ore minerals in the subsoil [3]. Vana-dium concentrations in igneous rocks are higher than inacidic and in siliceous rocks, whereas the V content of meta-morphic and sedimentary rocks is intermediate between thecontent of basic and acidic igneous rocks [3]. In soils, V ismainly associated with Fe(hydr)oxides, clay minerals, andorganic matter and it can also occur as discrete mineralphases such as carnotite (K2(UO2)2(VO4)2·3H2O) and vana-dinite (Pb5(VO4)3Cl) [2, 4, 5]. As an element of group VBand with atomic number 23 in the periodic table, it canacquire several oxidation states (+III, +IV, +V) [2, 6].

Beside its natural occurrence in soils and sediments, sev-eral anthropogenic sources can cause an enrichment or evenpollution of soils and sediments with V. The main anthro-pogenic sources of V are (1) the use of V as a catalyst in

the metal-, cement-, electronics, and textile industry; (2) theproduction of certain metals such as iron and uranium, withV as a secondary product [2], vanadium concentrations upto 738 mg/kg [7] and 3505 mg/kg [1] have been reported inmining areas; (3) air emission of V from the combustion offossil fuels [8]. The majority of the emissions of V into theair are taken up by the soil surface, especially when soils arerich in organic matter [9]. In Flanders (Belgium), the petro-chemical and chemical industry are the main anthropogenicsources for V. Other minor contributors to V emissions arethe food industry and the non-ferro industry [10].

Most food crops and also other plants contain traceamounts of V [6] because small concentrations of V stimu-late plant growth and act as a catalyst for nitrogen fixation[11]. When plants and crops are exposed to enhanced Vconcentrations in soil, they are usually smaller than the sameplants grown on noncontaminated soils [1, 2, 12]. V concen-trations in porewater above 3 mg L−1 result in importanttoxic effects for plants [3]. Kasai et al. [13] found that 0.1 μMvanadate had slight effects on seed germination, growth,photosynthesis, respiration, and ATPase activity of wheat andrye grass. The use of certain fertilizers ((NH4)2PO4, Na3PO4)


increases the solubility, mobility, and uptake of V by plants[14]. The oxidation state of V is primarily controlled by theEh (redox potential) and pH of a system. Under oxidizingand more alkaline conditions, vanadium is found in anionicform, as vanadate (HVO2−

4 or H2VO−4 , oxidation state +V),

whereas vanadyl (VO2+, oxidation state +IV) has cationicproperties and occurs under more acid conditions [3]. Com-plexes with fluoride, sulfate, or oxalate can increase the solu-bility of V under oxidizing conditions [15]. Acute poisoningof cattle by uptake of V-containing soil particles sticking tograss has been reported by McCrindle et al. [16].

Vanadium toxicity for human beings can vary widely, de-pending on its oxidation state. Clinical picture of poisoning(starting from 10 mg per day) shows different toxic effects ofvanadium on the respiratory, circulatory and central nervoussystems, the digestive organs, kidneys, and skin [17]. In thepast, V has also been used as a therapeutic for the treatmentof diabetes.

1.2. Alluvial Soils, Overbank Sediments, Floodplain Sediments.Darnley et al. [18] made a distinction between alluvial soilsaccording to the catchment basin size: floodplain and over-bank sediments are alluvial soils of large and small flood-plains, respectively, but both consist of fine-grained (silty-clay, clayey-silt) sediments. Floodplain and overbank sedi-ments are deposited during flood events in low energy envi-ronments [19]. Overbank sediments are considered a repre-sentative sampling medium for geochemical mapping [19].In areas of historical mining and pollution, only one over-bank sediment sample is not always representative for thegeochemical characteristics of a large catchment basin [20].De Vos et al. [4] evaluated the use of overbank sediments asa medium for geochemical mapping in industrialized areas.Samples of lower overbank sediment samples (i.e., samplestaken deeper in the overbank sediment profile) reflect thenatural, pristine situation, while upper overbank sedimentsare influenced by anthropogenic chemical contamination.From an environmental point of view, the use of overbanksediments is interesting because background concentrationsare an essential reference point to evaluate the pollutionstatus of soils and sediments.

1.3. Legal Framework with regard to V Contamination. In theenvironmental legislation of Flanders and Wallonia (whichare the two main regions in Belgium, each with an own envi-ronmental legislation), threshold vanadium concentrationsare defined for emissions in air, surface water, and groundwa-ter. Although background and intervention values are pro-vided in the Flemish and Walloon Soil Decree for As, Cd,Cr, Cu, Hg, Ni, Pb, Zn, and organic contaminants, no normvalues are provided for V. Nevertheless, several Europeancountries established norm values for V in soils (Table 1).

The norm values in Table 1 differ because of the use ofdifferent models, software, and criteria with regard to humantoxicology and ecotoxicology [21]. Additionally, natural(background) concentrations of trace elements in soils andsediments are dependent on soil and sediment properties(clay content, organic carbon content, etc.) and on the geo-

logical substratum, which makes it difficult to compare thesenorm values [23]. Moreover, total concentrations of traceelements do not give a good indication of the mobility andtoxicity of an element [2].

The aim of this study was (1) to evaluate the V-contentof alluvial soils (overbank sediments, floodplain sedimentsand dredged sediment-derived soils) along some Belgian andEuropean rivers and to compare these values with normvalues established in other European countries, (2) to es-tablish regression equations that allow to explain the Vcontent in alluvial soils based on major element compositionand/or physico-chemical soil properties such as clay- and or-ganic matter content, (3) to assess the actual and potentialrelease (“mobile” and “mobilizable” fractions resp.) of Vfrom these soils by means of standardized European extrac-tions and leaching tests (single extractions with CaCl2 andammonium-EDTA). It is not the purpose of this study topropose norm values for V in (alluvial) soils, but the regres-sion equations obtained in this study can be useful toestablish whether V concentrations that are measured in allu-vial soils can be considered as “normal” or not.

2. Methodology

2.1. Sampling Locations and Sampling

2.1.1. Belgian Samples. Between November 2006 and Octo-ber 2009, alluvial sediments, consisting of overbank sedi-ment- or dredged sediment-derived soils, were sampledalong 3 different rivers in Belgium (Figure 1).

In northern Belgium, overbank sediment- and dredgedsediment-derived soils were sampled in the catchment ofthe Scheldt river (catchment area in Belgium = 13336 km2),which is divided in 11 subcatchments, including the Leie andDemer catchments. The Leie river (north-western part ofBelgium) is an important tributary of the river Scheldt and itstotal catchment area, consisting of Tertiary sands and clays, isabout 3675 km2 (in Belgium). The Leie river is characterizedby a regular flooding regime. In 1976, a part of the river wasstraightened and dredged sediments were stored along theriver. Heavy metals in the dredged-sediment-derived soilsoriginate from industrial activities and agriculture. Mostof the dredged-sediment-derived soils are currently used asmeadows or forests. The Grote Beek (Demer catchment,1272 km2) river (northern Belgium) is underlain by glau-conite-rich medium-grained sands (Diestian Formation).Organic- and iron-rich wetland soils have developed alongthis stream. The river follows a very meandering path and asa consequence several flooding zones occur along the river,which are flooded a few times a year during periods of heavyrainfall. The major source of pollution is a factory that isproducing phosphate fertilizers by the refinement of phos-phate ores. The majority of the alluvial soils along the GroteBeek consist of unused (i.e., not used for living, industry oragriculture) land, situated in a natural reserve.

In eastern Belgium and in the southern part of TheNetherlands, overbank sediments were sampled along theGeul river (Figure 1). The Geul river, a tributary of the


Table 1: Norm values for vanadium (total concentration in mg kg−1) in soil in European countries [21, 22].

CountryNegligible Warning Unacceptable Unacceptable

risc risc risc (residential) risc (industrial)

The Netherlands 42 — 250 —

Czech Republic 180 340 450 —

Slovenia 120 200 500 —

Finland — 100 150 250

Sweden — 200 — —

Italy — — 90 250

Lithuania — — 150 —

Russia — — 150 —

Leie

Scheld

tLeie

Scheldt

Demer

Meu

se

Diest

Brussels

Antwerpen

Kortrijk

Kortrijk

Ghent

Ghent

GeulSippe-

naeken

Plombieres

Epen

BelgiumThe Netherlands

m

Scheldt

km

km

0

0 050

10

10

00

N

D

m

Grote Beek

Zwart WaterHulpe

Scherpenheuvel(Zichem)

(Molenstede Diest)

(Deurne Diest)

Nete

NL

F

L

2000

Zenn

e

0

Figure 1: Overview of the Belgian sampling locations.

Meuse, flows in sandstones (Upper Devonian) and carbon-ates and shales (Lower and Middle Carbonian) [24]. Thedischarge of the Geul river depends largely on the amount ofrainfall and is characterized by a flashy regime with a se-quence of small floods in autumn and winter and a few ma-jor floods in summer [25]. From the Middle Ages until thebeginning of the 20th century, extensive Zn-Pb mining andsmelting was carried out in Plombieres and La Calamine(eastern Belgium). Besides the important amount of wastethat is stored in the large mine tailing piles, overbank sedi-ments along the nearby Geul river are severely contaminatedwith Zn, Pb, and Cd [24]. The Geul drains a predominantlyrural catchment with a relatively low population density.

At each location, vertical profiles were dug and samples(1 kg of material per sample) were taken at depth every 5 to

20 cm, depending on visual differences in color, organic mat-ter content and/or texture, yielding 3 to 5 samples per hori-zon. In total, 206 samples were taken. Along the Leie river,samples were taken in dredged sediment-derived soils. Alongthe Geul, profiles were sampled within overbank sediment-derived soils and along the Grote Beek both overbanksediment- and dredged sediment-derived soils were sampled.

2.1.2. European Samples. “The FOREGS Geochemical Base-line Mapping Programme’s main aim is to provide high-quality, multipurpose environmental geochemical baselinedata for Europe. The need for this type of data was justifiedby the first Working Group on Regional GeochemicalMapping immediately after the Chernobyl accident in 1986,when it was realized that a baseline for radioactive and other


Table 2: Comparison between certified and measured (average of 3 replications) values for V in 3 certified reference materials.

Certified value (mg kg−1) Measured value (mg kg−1)

SRM2710 Montana soil 76.6 69.8 ± 4.7

GBW07311 Soil 47 ± 5 42.6 ± 3.3

GBW07411 Stream sediment 88.5 ± 6.9 84.1 ± 6.8

polluting elements could not be defined” [26–28]. After sev-eral attempts to compile existing regional geochemical data-bases in Europe, it was clear that the establishment of aharmonized European wide geochemical database was essen-tial. The Forum of European Geological Surveys’ Directors(FOREGS) approved a Geochemical Baseline Mapping Pro-gramme in 1996, and in 1997 an agreement on the prin-ciples of field and analytical methodologies was obtainedand the field methodology was tested, finally resulting in the“FOREGS Geochemical Field Manual,” that was published in1998 [29]. In the present study, data from floodplain sedi-ments samples in 26 countries that were involved in theFOREGS Geochemical Baseline Mapping Programme areused.

From each sampling site, 2 kg of top floodplain sediment(sampling depth 0–25 cm) was collected. The samplingprocedure for floodplain soils is described in the field manual[29].

2.2. Physicochemical Soil Characterization

2.2.1. Belgian Alluvial Soils

Physicochemical Characterization. All analyses were per-formed on air-dried soil samples. For the determination ofelement concentrations and organic carbon content, partof the soil sample was disaggregated in a porcelain mortarand sieved (<1 mm). pH(H2O) was measured in a soil/watersuspension (1/2.5). Organic carbon was determined accord-ing to the Walkley and Black method [30]. Grain size wasdetermined by laser diffraction analyzis (Malvern MastersizerS long bed). Total element concentrations (Al, As, Ba, Cd,Cu, Cr, Co, Ca, Fe, K, Mg, Mn, Ni, Pb, V, and Zn) weredetermined in all the samples, but the discussion will mainlyfocus on V. One gram of sample was dissolved in 4 mLHClconc, 2 mL HNO3conc, and 2 mL HFconc in a Teflon beaker.The mixture was gently heated on a hot plate until halfdry, subsequently reattacked with the same amount of thethree acids and heated until completely dry. The residuewas redissolved with 20 mL 2.5 mol L−1 HCl and filtered(Whatman 45). Finally, the solution was diluted to 50 mLwith distilled water. This destruction method was chosenbecause we have a lot of experience with this method in ourlaboratory and it allows to compare the results with previousstudies performed in our laboratory. Concentrations ofmajor elements (Fe, Al, Ca, K, and Mg) were measured withFlame atomic absorption spectrometry (FAAS, Varian AA6),whereas an Induced coupled plasma mass spectrometry(ICP-MS, HP 4500 Series) was used for As, Ba, Cd, Cu,Cr, Co, Mn, Ni, Pb, Vm and Zn. Three certified referencematerials, namely, GBW07411 Soil, (a Chinese soil),

GBW07411 sediment (a Chinese stream sediment), and SRM2710 Montana soil (a soil collected from the upper 10 cm ofpasture along a creek and contaminated by overbank dep-osition of contaminated sediments) were analyzed for qualitycontrol. Total certified V concentrations in these materialsare determined by XRF and/or neutron activation analysis,whereas a dissolution with 3 strong acids (HCl, HNO3, andHF) is used in the present study, followed by a measurementwith ICP-MS. Despite the use of hydrofluoric acid, a smallresidue (usually less than 5% of the dry mass of the sample)is not dissolved. A previous investigation in our laboratoryshowed that this residue consists of quartz particles andsometimes also phyllosilicates. This possibly explains whymeasured V concentrations are slightly lower compared tocertified values (Table 2). Nevertheless, the use of a disso-lution with strong acids and ICP-MS determination allowsa better comparison with data from the FOREGS project,where V has been determined by ICP-MS after aqua regiadissolution and by XRF. It is clear that the procedures forthe determination of total element concentrations in soilsand sediments should be well established, especially in large-scale studies, since this is essential to compare results ofdifferent studies and/or laboratories. Additionally, triplicateanalysis was performed for some samples. Relative standarddeviations on triplicate analysis (3 separate subsamples) werebelow 5% for all elements, except Ca and Al (below 10%).

Single Extractions. Single extractions were performed on65 samples of alluvial soils from the Leie and Grote Beekrivers. For the CaCl2 and ammonium-EDTA extraction, theprotocol of the SMT (Standards Measurement and Testing)program [31] was followed. 20 mL of a 0.01 mol L−1 CaCl2 or20 mL of a 0.05 mol L−1 ammonium-EDTA solution, respec-tively, was added to 2 g of air-dried sediment in a centrifugetube. The suspension was shaken for 3 h or 1 h, respectively,in a reciprocal shaker, centrifuged (3500 rpm, 10 minutes),decanted, and filtered (0.45 μm). After measurement of pH,the CaCl2 extract was acidified with concentrated HNO3 tobring the pH to <2. The ammonium-EDTA extracts were notacidified prior to analysis to prevent precipitation of EDTAsalts at very low pH. A reference material (CRM 483) certifiedfor its ammonium-EDTA and CaCl2- extractable content ofCd, Zn, Cu, and Pb was also included. No certified valuesare provided for V, but the data obtained in this study arepresented in Table 3 in order to allow comparison with otherstudies in the future. Standard deviation between triplicateextractions was less than 5%. The data of the ammonium-EDTA extracts are presented in mg/kg dry matter, sinceEDTA-extractable element concentrations are often com-pared to total element concentrations. The V concentrationsin the CaCl2 extracts are expressed in μg L−1, to facilitate the


Table 3: Values for V measured in the CaCl2 (in μg L−1) and ammonium-EDTA (in mg kg−1) extracts performed on certified referencematerial CRM 483 according to the BCR extraction protocol for single extractions with CaCl2 0.01 mol L−1 and ammonium-EDTA0.05 mol L−1.

CaCl2 0.01 mol L−1 ammionum-EDTA 0.05 mol L−1 Total concentration

μg L−1 mg kg−1 mg kg−1

CRM483 5.7 ± 0.9 4.3 ± 0.1 64.5 ± 3.7

comparison with, for example, porewater concentrations ofV.

2.2.2. Alluvial Soils from the FOREGS Database. Detailsabout the analytical methods are provided in Sandstrom etal. [32]: total concentrations of MgO, P2O5, K2O, CaO, TiO2,V, Cr, MnO, Cs, Ba, La, and Ce were determined using energydispersive polarized X-ray fluorescence spectrometry. In thefloodplain soil samples, V was also determined by ICP-MSafter aqua regia destruction. Because the data in both datasetsare not always in the same order and data are sometimesmissing, a comparison between data obtained with bothmethods was not performed. In the present study, mainly thedata from the XRF analysis will be used.

2.3. Statistical Analysis. Statistical analysis was performedwith the software package SPSS 16.0 for Windows.

Descriptive statistics (average, median, minimum, maxi-mum, standard deviation, and variance) were calculated foreach variable. The 90th percentile of element concentrationsin the upper soil layer of noncontaminated soils is oftenconsidered representative for the background value in soils[22, 33]. Because our datasets contained both contaminatedand noncontaminated samples, no attempt was made to pro-pose background values. By comparing average and medianvalues, soils with enrichment in V can be detected [33].

The normal distribution of the variables was checked bymeans of the Kolmogorof-Smirnov (K-S) test and correla-tions between variables were tested by calculating two-tailedPearson correlation coefficients for the log transformed val-ues. Analysis of variance (ANOVA) was applied to compareaverages of the variables between river basins. Multiple linearregression according to the stepwise method was performedto deduce possible causal relationships between the variables.Attention was mainly paid to the possibility of predictingtrace element concentrations in sediments based on majorelement composition, pH, clay, and organic matter content.Different assumptions of the linear regression (normalityof the de residues, autocorrelation, quasi-multicollinearity(QMC), and heteroscedasticity) were tested.

For the statistical analysis, the recommendations of Web-ster [34, 35] (reporting mean values with standard errors,performing linear regressions, etc.) were taken into account.

3. Results and Discussion

3.1. Total Concentrations of V in Floodplain Soils. Averagetotal V concentration in the 3 Belgian catchments and inthe European FOREGS database were in the range 45–87 mg kg−1. The high average V concentrations in the alluvial

soils of the Grote Beek catchment is mainly due to 7 outli-ers with V concentrations between 189 and 301 mg kg−1.Alluvial soils of the Geul river are characterized by significantlower V concentrations compared to the Leie and Grote Beekrivers. V concentrations in floodplain soils are rarely report-ed in literature. Peh and Miko [36] measured average Vconcentrations of 70 mg kg−1 in Croatia, which is in line withthe values reported here. Ivanov and Kashin [37] recorded(AAS-measurement after aqua regia destruction) average Vconcentrations of 110 mg kg−1 in alluvial deposits in Trans-baikalia (Russia), and Wigginton and Price [38] found (ICP-OES determination after aqua regia destruction) concentra-tions between 8 and 17 mg kg−1 in floodplain soils of BayouCreek (UK). On the 4th October 2010, red mud suspensionwere released from the Ajkai Timfoldgyar Zrt alumina plant.Among other trace elements, mud samples at the sourcecontained important V concentrations (>1000 mg kg−1), andfloodplain sediments downstream were significantly affectedby the red mud [39].

Total V concentrations in floodplain soils from theFOREGS database are represented on Figure 2. Since the datapresented on Figure 2 are characterized by a right kurtosisand thus not normal distribution, concentrations represent-ed on the map are strongly influenced by locally enhanced Vconcentrations [40]. Data for the Belgian catchments are notpresented on a map, since the data originate from 3 distinctlocations, only covering a limited area in Belgium.

Data observations which lied more than 1.5 interquartilerange lower than the first quartile or 1.5 interquartile rangehigher than the third quartile were considered as an outlier.In the Belgian data for V, 7 outliers were observed, between189–301 mg kg−1, all coming from samples from the alluvialsoil samples of the Grote Beek. De Grote Beek is also char-acterized by the highest average V concentrations and animportant difference between average and median values forV, which indicates an anthropogenic source for V [33]. In thisarea, phosphate ores are processed by a plant that is emit-ting its waste water directly into the river. V is known asa secondary product of phosphate production [8]. In accor-dance to this, the correlation coefficient between V and P inalluvial soils is also very high in the samples from the GroteBeek river.

When the V concentrations in these samples are com-pared with the norm values in Table 1, 11 samples of theGrote Beek alluvial plain exceed the value for “unacceptablerisk” in residential areas established in Finland, Lithuania,and Russia (150 mg kg−1), whereas 3 samples exceed theDutch “negligible risk” value of 250 mg kg−1. The samplesfrom the catchment of the Leie river are contaminated withheavy metals and often characterized by an elevated clay-


Geoch

emicalbase

line

mappin

g

XRF, detection limit 2mg kg −1

Number of samples 747 Median

0

1724.532.641.250.159.469.17989.399.7110

120

130140

270

11

22

10

10

34

100

44

270

56

69

0

81

94

20

120

500 1000

V

V

V

Vanadium

(Kilometers)

0.1

1

51025

50

75

9095

99

99.9

floodplain

Freq

uen

cy (

%)

V mg kg−1

Carnary Islands

<2

56 mg kg−1C

um

ula

tive

freq

uen

cy (

%)

Figure 2: Total V concentrations in alluvial soils samples in 26 European countries [27].


Table 4: Summary statistics for total V concentrations (in mg kg−1), based on Belgian (this study) and European [28] data. Standarddeviation is reported within brackets. n: number of samples.

dataset area mean median min max n

Belgian samples

3 Belgian floodplains 69 (43) 53 18 301 206

floodplain of the Geul 45 (6) 46 30 57 75

floodplain of the Grote Beek 87 (55) 66 28 301 94

floodplain of the Leie 71 (28) 69 18 160 37

FOREGSEurope (XRF) 59 (35) 56 1 266 749

Europe (aqua regia) 56 747

and organic matter content. Only one sample from alluvialplain of the Leie river is characterized by a V content above150 mg kg−1 (total V concentration is 161 mg kg−1).

The samples from the Geul-catchment, finally, are char-acterized by relatively low total V concentrations (average45 mg kg−1, Table 4). The average V concentration in thealluvial soils of the Geul river is below the 90th percentileof the European dataset, which could be considered as thebackground value [33], as well as below the average V con-centration in European alluvial soils. The maximal V concen-trations is 57 mg kg−1 (Table 4) and thus below all thresholdvalues established in Table 1, except the value for “negligiblerisk” from The Netherlands (42 mg kg−1). The average Vconcentration in the soils along the Geul is very close tothe median concentration, which indicates that V is not ofanthropogenic origin [33]. The subsoil in the Geul catch-ment consists of sand and shale and other studies alsoshowed that total V concentrations in soils developed onshale are lower compared to more clay-rich substrates [41].

In the European FOREGS dataset, “the spatial distri-bution of V is clearly related to the bedrock geology andmineralization, especially mafic and ultramafic lithology,and also clay-rich soil with high Al2O3 content” [15]. Thehighest V concentration is found in the mineralized Oslorift (266 mg kg−1). In England and Germany, elevated V con-centrations (224 mg kg−1) are respectively due to industrialpollution and coal and oil combustion [15].

3.2. Relation of V with Other Elements

3.2.1. Correlations. Because the majority of the data (traceelement and major element concentrations) showed a rightkurtosis and were significant on the K-S test, the data werelog transformed. Concerning the FOREGS data, the corre-lation coefficients in Table 5 were calculated with the datafrom the XRF determination, since more elements that arerelevant for this paper were measured with this method com-pared with the aqua regia method. Unfortunately, it was notpossible to merge the results of the aqua regia and XRF deter-mination into one table, because the data were not in thesame order and there were several missing samples for eachof the methods. This inconvenience of the FOREGS data-base has also been mentioned by [40].

In general, Mg, Cr, and Al showed the strongest cor-relation with V (Table 5). Among the trace elements, Cr ischaracterized by a strongly positive correlation with V in thesamples of the Grote Beek (R = 0.933). Other studies also

demonstrated a high correlation between Cr and V in soilsand sediments [23, 42]. Cr and V can substitute for eachother in the crystal structure of clay minerals and vanadateand chromate show a similar behavior in soils.

In the catchment of the Grote Beek, elevated Fe concen-trations were generally found, with a maximal concentrationof 7.66% Fe, which could be explained by the occurrence ofiron-bearing sands and sandstone as well as glauconite in thesubsoil [43].

Fe and Al also show a positive correlation, which is oftenobserved in soils and sediments (e.g., [15, 42, 44–46]).

Data for the clay and organic carbon content were onlyavailable for the floodplain soils of the Leie. Vanadium showsa significant positive correlation with the clay content(0.713), which is in accordance with the positive correlationwith Al and other clay-associated elements and with the or-ganic carbon content (0.657). Although the association of Vwith organic material is also shown in other studies [2, 46],some authors [42] report a significant negative correlationbetween V and organic matter content. Połedniok and Buhl[5] concluded that the amount of V bound to organicmaterial is lower in industrial areas compared to rural areas.

For the XRF-data from the FOREGS database, vanadiumin the European floodplain soil samples is strongly positivelycorrelated with Fe and Ti, and to a lesser extent with Al.

These positive correlations can be explained by the factthat (hydr)oxides of Fe, Ti and Al are good adsorbers for Vions [47, 48] and because of the occurrence of these elementsin clay minerals. Moreover, the atomic radius of the V ion issimilar to the radius of Fe and Al [2, 15].

3.2.2. Regression Equations to Predict V Concentrations inAlluvial Soils. Multiple linear regressions were performedwith Fe, Mg, Ca, K, Al, P, and the clay (<2 μm fraction),and organic matter content as independent variables. Clayand organic matter were included because they are alreadyused in the standardization of background values for heavymetals in soils and sediments [49]. Fe is a major componentof Fe(hydr)oxides, and is also a constituent of sheet silicates.In soils and sediments, Ca and Mg are dominantly foundin sheet silicates when the parent material is Mg-rich (e.g.,ultrabasic rocks) and in carbonate minerals (e.g., CaCO3) innonacidic soils.

Regression equations were constructed according to thestepwise method [50]. In order to only include the mostsignificant independent variables. With exception of the soilsfrom the Grote Beek, all regression equations in Table 6 are


Table 5: Two-tailed Pearson correlation coefficients of V with other elements measured in floodplain soils of Belgian and European rivers.

Belgian data Grote Beek Geul Leie FOREGS

Mg 0.793∗∗ 0.821∗∗ 0.391∗∗ 0.778∗∗ 0.570∗∗

Al 0.030 0.785∗∗ 0.686∗∗ 0.870∗ 0.680∗∗

P 0.626∗∗ 0.556∗∗ −0.028 0.280 0.374∗∗

K 0.681∗∗ 0.343∗∗ — 0.605∗∗ 0.433∗∗

Ca 0.481∗∗ 0.088 −0.100 0.170 0.064

Cr 0.729∗∗ 0.933∗∗ 0.177 0.631∗∗ 0.665∗∗

Mn −0.079 0.516∗∗ 0.211 0.397∗ 0.640∗∗

Fe 0.668∗∗ −0.118 0.461∗∗ 0.827∗∗ 0.896∗∗

Co 0.201∗∗ 0.825∗∗ 0.399∗∗ 0.672∗∗ 0.841∗∗

Ni 0.424∗∗ 0.551∗∗ 0.322∗∗ 0.717∗∗ 0.706∗∗

Cu 0.513∗∗ 0.651∗∗ 0.235∗ 0.484∗∗ 0.649∗∗

Zn 0.021 0.665∗∗ 0.152 0.382∗ 0.556∗∗

As 0.575∗∗ 0.411∗∗ 0.249∗ 0.594∗∗ —

Se 0.468∗∗ 0.365∗ 0.184 0.715∗∗ —

Sr 0.667∗∗ 0.548∗∗ 0.601∗∗ 0.375∗ —

Cd 0.428∗∗ 0.314∗∗ 0.121 0.365∗ —

Ba 0.439∗∗ 0.437∗∗ 0.561∗∗ 0.628∗∗ 0.463∗∗

Pb −0.128 0.615∗∗ 0.211 0.461∗∗ 0.347∗∗

Ti — — — — 0.869∗∗

OM — — — 0.657∗∗ —

Clay — — — 0.713∗∗ —

Loam — — — 0.054 —

Sand — — — −0.511∗∗ —

S — — — 0.463∗ —∗

: significant at α = 0.05 (two-tailed).∗∗: significant at α = 0.01 (two-tailed).

characterized by one or more outliers. Omission of the out-liers from the dataset resulted in a better fit for the regressionequations (Table 6). Additionally, the QMC, homoscedas-ticity, normality of the residues in independent residuesimproved in most of the cases.

In general, total V concentrations could be predicted verywell by the independent variables Al, Fe, and Mg (Table 6).An example of a good fit between measured and predictedvalues is presented in Figure 3 for the alluvial soils of the Leie.

For the Belgian dataset, Al was an important explainingvariable, together with Fe, Mg, and/or P (Table 3). Partic-ularly for the Grote Beek alluvial soils, P is an explainingvariable in the regression equation for V, which can be relatedto the emissions from the phosphate ore processing plant.

Despite the fact that major elements such as Fe, and Mgare also a component of sheet silicates, they are also majorcomponents of resp. Fe-(hydr)oxides and carbonates. More-over, the attribution of the term “clay fraction” to the <2 μmfraction is sometimes misleading. Beside clay minerals, Fe-rich minerals such as Fe-oxi/hydroxides and phyllosilicatesare often concentrated in the clay fraction (<2 μm) [51]. Ageochemical approach using the total concentrations of con-servative elements may be used to overcome these difficulties(e.g., [52]). Before processing the dataset, the potentialanthropogenic input of the element and diagenetic processesthat may alter its concentrations in sediments should bechecked [53]. For the European data, Fe, Al, and Mg werethe most significant explaining variables (Table 6).

y = 0.824x + 0.3424

R2 = 0.824

1.2

1.4

1.6

2

2.2

2.4

1.21.4 1.6

1.8

1.8 2 2.2 2.4

Measured (logV)

Pre

dict

ed (

log

V)

Figure 3: Predicted versus measured V concentrations in thefloodplain soils of the Leie river.

3.3. Single Extractions

3.3.1. Prediction of “Mobile” Metal Concentrations. The com-position of soil porewater is important from an environ-mental point of view because it gives an indication of the“actual mobility” of heavy metals and because the uptake oftrace elements by plants occurs via the porewater. Moreover,porewater is also the carrier for elements to the groundwater.


Table 6: Regression equations for the different areas with V (logtransformed) as dependent variable.

Area Regression equation R2

3 Belgianfloodplains

alog V = −3.172 + 0.433 log Al + 0.367 log Mg + 0.399 log Fe + 0.058 log P∗∗ 0.831blog V = −2.956 + 0.424 log Al + 0.341 log Mg + 0.312 log Fe + 0.068 log P∗∗ 0.920

floodplain of theGeul

alog V = −1.424 + 0.507 log Al + 0.174 log Fe∗∗ 0.606blog V = −1.342 + 0.503 log Al + 0.160 log Fe∗∗ 0.634

floodplain of theGrote Beek

alog V = −2.820 + 0.761 log Al + 0.196 log Mg +0.207 log P ∗∗ 0.809

floodplain of theLeie

alog V = −4.244 + 0.945 log Al + 0.368 log Mg + 0.315 log clay∗∗ 0.824blog V = −3.870 + 0.775 log Al + 0.472 log Mg + 0.330 log clay∗∗ 0.900

Europealog V = 1.041 + 1.32 log Fe + 0.413 log Al + 0.102 log Mg∗∗ 0.836blog V = 1.108 + 1.33 log Fe + 0.279 log Al + 0.086 log Mg∗∗ 0.880

aIn the first regression equation, outliers are included.b The second regression equation is without outliers. ∗∗Significant at 0.01 level.

The composition of the CaCl2 extract is often consideredrepresentative for porewater composition. Contaminants inthe porewater can move to deeper soil layers or to thegroundwater by infiltrating water. In the present study, aver-age CaCl2-extractable V concentrations were 2 μg L−1 for thesamples from the Leie and 4 μg L−1 for the samples from theGrote Beek (Table 7), which is far below the value of3 mg L−1, which is proposed by Edwards et al. [3] and wouldcause significant toxic effects to plants.

Metal partitioning in soils can be quantified by models inwhich metal concentrations in the porewater are described asa function of the metal binding solid phases such as Fe- andAl-(hydr)oxides, organic matter, and clay and as a functionof soil characteristics that influence heavy metal partitioning,such as pH. The ratio between total metal content bound toa soil relative to its concentration in the soil solution isoften represented by Kd coefficients. However, such a mod-el assumes that the sorption capacity of a material is inde-pendent of the soil properties (organic matter content, pH,clay content, etc.) and therefore, single Kd values are not ap-propriate to predict metal solubility in soil. Therefore, severalauthors indicated that metal solubility could be predictedfrom soil properties. For example, Houba et al. [54] observeda significant relationship between the distribution coefficient(Kd) of Cd in soils and the pH of the CaCl2 extract. Accordingto McBride et al. [55] most of the variability of metal solu-bility in soil is explained by pH, organic matter content andtotal metal concentrations.

A semimechanistic approach was developed by Sauveet al. [56] which is based on the assumption that exchange-able metals and protons compete for adsorption on soil-ex-change sites. In this model, pH, total metal concentrations,and soil organic matter are used to predict dissolved metalconcentrations.

In the present study, a modified version of this competi-tive adsorption model was applied, since ammonium-EDTA-extractable metal concentrations were used instead of totalmetal concentrations. EDTA extractions are often used toestimate the potentially “available pool” of metals (i.e., thepool that can deliver metals from the solid phase of the soilto the soil solution in a relatively short time period). Stepwise

multiple linear regressions was performed with pH(CaCl2),organic carbon content, clay content, CaCl2-extractable V inP concentrations, dissolved organic carbon (DOC) contentin the CaCl2-extracts, and ammonium-EDTA-extractable Vin P concentrations. Phosphorus was included because phos-phate can influence the release of V from soils [7, 13, 57].

The distribution of an element between the liquid andsolid phase depends strongly on the speciation of the selec-ted element. Because vanadium is known to occur in mul-tiple oxidation states, the actual distribution of the oxidationstates was first calculated using the speciation code Visual-MinteQ. According to these calculations, the main vanadiumspecies encountered in the CaCl2 extracts is HVO2−

4 , whichmeans that V occurs as an anion that will have the tendencyto be desorbed when pH rises.

For the Grote Beek floodplain soils, only the pH of theCaCl2 extract and ammonium-EDTA-extractable V concen-trations were retained in the final regression model, as theother variables did not contribute to an improvement of theregression equation. For the Grote Beek river catchment,“mobile” V concentrations in floodplain soils could be pre-dicted using the following fitting equation:

log [V]s = 5.356− 1.041 pH + 1.338 log [V]a

R2 = 0.840(1)

with [V]s = the V concentration in the CaCl2 extract(mg L−1), [V]a = the “available pool” of V (mg kg−1).

For the floodplain soils along the Leie, the pH of theCaCl2 extract, EDTA-extractable V-concentrations and or-ganic carbon content were retained in the final regressionmodel, resulting in the following regression equation:

log [V]s = 1.313− 1.969 log clay + 0.378 log [P]a

+ 1.150 log OC

R2 = 0.590

(2)


Table 7: Summary statistics for V concentrations in the CaCl2 (in μg L−1) and ammonium-EDTA (in mg kg−1) extracts. Standard deviationis reported within brackets. n: number of samples.

mean median min max n

Leie river

CaCl2 (μg L−1) 2.0 (1.7) 1.8 <0.1 6.7 39

ammonium-EDTA (mg kg−1) 2.4 (2.2) 2.0 <0.1 9.8 39

Grote Beek river

CaCl2 (μg L−1) 4.4 (4.5) 2.0 <0.1 12.2 27

ammonium-EDTA (mg kg−1) 11.2 (12.1) 5.0 1.3 35.3 27

with [V]s = the V concentration in the CaCl2 extract(mg L−1), [P]a = the “available pool” of P (mg kg−1), OC= the organic carbon content (%), clay = the clay content(in %).

3.3.2. “Potential Availability” of V. Ammonium-EDTA wasalso used to estimate the “mobilizable” metal concentrations.EDTA is a nonselective reagent that exhibits a strong capacityto complex metals. EDTA was shown to dissolve carbonates,thereby mobilizing occluded elements [58]. Borggaard [59]showed that EDTA extracts amorphous Fe-oxides, but thisdissolution is very slow in the presence of other metal-chelate complexes [60]. It is also able to form organometalcomplexes, which compete with organic matter in soil. Addi-tionally, complexes of EDTA with heavy metals such as Cd,Cu, Pb, Ni, and Zn are more stable than complexes betweenV and EDTA, which results in a lower ammonium-EDTA-extractable fraction of V when the heavy metals mentionedbefore are present in high concentrations [61]. The averageEDTA-extractable V concentration in the floodplain soils ofthe Leie was 2.46 mg kg−1 (Table 7) which is comparable withaverage EDTA-extractable V concentrations (3.40 mg kg−1)reported by Gabler et al. [61]. In the samples from theGrote Beek river, the average ammonium-EDTA-extractableV concentration was 11.2 mg/kg. Floodplain soil samplesalong the Grote Beek are contaminated with V and displayelevated V concentrations, which can also explain for thehigher ‘potentially mobile V-content in these soils. In relativeconcentrations, however, only 3.5% and 9.8% of the total Vcontent in the floodplain soil samples of, respectively, theLeie and the Grote Beek were extracted with ammonium-EDTA, pointing to the fact that the majority of the V in thesamples is characterized by a very low mobility. In the highlycontaminated samples of the Grote Beek, a more detailedstudy of V mobility may nevertheless be useful to elucidatethe potential release of V under changing environmentalconditions (changes in pH, Eh, etc.).

4. Conclusion

Average V concentrations in Belgian alluvial soils(69 mg kg−1) are in general comparable with the average Vcontent in European alluvial soils (56–59 mg kg−1). In one ofthe river catchments analyzed in this study, a contaminationof alluvial soils, originating from a phosphate ore processingplant, was detected. By analysis of both the Belgian and Euro-pean data, the relationship between concentrations of major

elements and V was quantified. Fe, Al, and Mg were the mostsignificant variables that could predict the total V content inthe alluvial soils and the obtained regression equation couldbe used to predict V concentrations in alluvial soils. For thealluvial soils contaminated by the emissions from the phos-phate ore treatment plant, P was also an explaining variable.“Mobile” V concentrations, as estimated by the amount of Vreleased by a single extraction with CaCl2 0.01 mol L−1, werelow, even in the most contaminated soil samples. Despite thelow actual mobility of V, it might nevertheless be useful tostudy the effect of changing environmental conditions suchas soil acidification and fluctuating redox conditions (e.g., byinundation of the floodplain soils). This will be the subjectof a followup study, in order to obtain a better insight in thegeochemistry of V in soils and sediments.

Finally, it is clear that the procedures for the determina-tion of total element concentrations in soils and sedimentsshould be well established, especially in large-scale studies,since this is essential to compare results of different studiesand/or laboratories.

References

[1] N. Panichev, K. Mandiwana, D. Moema, R. Molatlhegi, andP. Ngobeni, “Distribution of vanadium(V) species betweensoil and plants in the vicinity of vanadium mine,” Journal ofHazardous Materials, vol. 137, no. 2, pp. 649–653, 2006.

[2] R. D. Schuiling, R. J. van Enk, and H. L. T. en Bergsma,“Natuurlijk voorkomen, mobiliteit en industrieel gebruik van“exoten” voorkomend in de Nederlandse bodem (Br, I, Ba, Sb,V, Sn, Co, Mo, Se),” 2003, http://www.geochem.nl/.

[3] R. Edwards, N. W. Lepp, and K. C. Jones, “Other less abundantelements of potential significance,” in Heavy Metals in Soils, B.J. Alloway, Ed., pp. 307–352, Blackie Academic/Chapman andHall, London, UK, 2nd edition, 1995.

[4] W. De Vos, J. Ebbing, R. Hindel, J. Schalich, R. Swennen,and I. Van Keer, “Geochemical mapping based on overbanksediments in the heavily industrialised border area of Belgium,Germany and the Netherlands,” Journal of Geochemical Explo-ration, vol. 56, no. 2, pp. 91–104, 1996.

[5] J. Połedniok and F. Buhl, “Speciation of vanadium in soil,”Talanta, vol. 59, no. 1, pp. 1–8, 2003.

[6] D. G. Barceloux, “Vanadium,” Clinical Toxicology, vol. 37, no.2, pp. 265–278, 1999.

[7] Y. Teng, X. Jiao, J. Wang, W. Xu, and J. Yang, “Environmentallygeochemical characteristics of vanadium in the topsoil in thePanzhihua mining area, Sichuan Province, China,” ChineseJournal of Geochemistry, vol. 28, no. 1, pp. 105–111, 2009.


[8] World Health organization (WHO), Vanadium Pentoxide andOther Inorganic Vanadium Compounds, WHO, Geneve, Switz-erland, 2001.

[9] B. J. Alloway, Heavy Metals in Soils, Chapman and Hall, Lon-don, UK, 1995.

[10] P. Lodewijks, C. Polders, K. Briffaerts, and H. Van Rompaey,“Evaluatie emissiereductiepotentieel voor diverse polluente-missies naar de lucht van de chemische sector, deel II, Ein-drapport,” Evaluation of the potential reduction of industrialemissions from the chemical sector, Part 2, Final report II.Study ordered by Aminal, Environmental Policy section,2004/IMS/R/186, August 2004.

[11] J. R. Benemann, C. E. McKenna, R. F. Lie, T. G. Traylor, andM. D. Kamen, “The vanadium effect in nitrogen fixation byazotobacter,” Biochimica et Biophysica Acta, vol. 264, no. 1, pp.25–38, 1972.

[12] A. Olness, R. Gesch, F. Forcella, D. Archer, and J. Rinke,“Importance of vanadium and nutrient ionic ratios on the de-velopment of hydroponically grown cuphea,” Industrial Cropsand Products, vol. 21, no. 2, pp. 165–171, 2005.

[13] M. Kasai, J. Yamazaki, M. Kikuchi, M. Iwaya, and S. Sawada,“Concentration of vanadium in soil water and its effect ongrowth and metabolism of rye and wheat plants,” Communi-cations in Soil Science and Plant Analysis, vol. 30, no. 7-8, pp.971–982, 1999.

[14] K. L. Mandiwana and N. Panichev, “The leaching of vana-dium(V) in soil due to the presence of atmospheric carbondioxide and ammonia,” Journal of Hazardous Materials, vol.170, no. 2-3, pp. 1260–1263, 2009.

[15] W. De Vos and T. Tarvainen, Eds., Geochemical Atlas of EuropePart 2: Interpretation of Geochemical Mops, Additional Tables,Figures, Maps and Related Publications, EuroGeosurveys &Foregs, Espoo, Finland, 2006.

[16] C. M. E. McCrindle, E. Mokantla, and N. Duncan, “Peracutevanadium toxicity in cattle grazing near a vanadium mine,”Journal of Environmental Monitoring, vol. 3, no. 6, pp. 580–582, 2001.

[17] B. Mukherjee, B. Patra, S. Mahapatra, P. Banerjee, A. Tiwari,and M. Chatterjee, “Vanadium—an element of atypical bio-logical significance,” Toxicology Letters, vol. 150, no. 2, pp. 135–143, 2004.

[18] A. G. Darnley, A. Bjorklund, B. Bolviken et al., A Global geo-chemical database for environmental and resource management.Final report of IGCP Project 259, Earth Sciences 19, UNESCOPublishing, Paris, France, 1995.

[19] R. T. Ottesen, J. Bogen, B. Bølviken, and T. Volden, “Overbanksediment: a representative sample medium for regional geo-chemical mapping,” Journal of Geochemical Exploration, vol.32, no. 1–3, pp. 257–277, 1989.

[20] J. Ridgway, D. M. A. Flight, B. Martiny, A. Gomez-Caballero,C. Macias-Romo, and K. Greally, “Overbank sediments fromcentral Mexico: An evaluation of their use in regional geo-chemical mapping and in studies of contamination frommodern and historical mining,” Applied Geochemistry, vol. 10,no. 1, pp. 97–109, 1995.

[21] J. Provoost, C. Cornelis, and F. Swartjes, “Comparison of soilclean-up standards for trace elements between countries: whydo they differ?” Journal of Soils and Sediments, vol. 6, no. 3, pp.173–181, 2006.

[22] C. Carlon, Ed., Derivation methods of soil screening values inEurope. A review and evaluation of national procedures towardsharmonization, European Commission, Joint Research Centre,Ispra, Italy, 2007.

[23] D. J. Spurgeon, P. Rowland, G. Ainsworth, P. Rothery, S. Long,and H. I. J. Black, “Geographical and pedological drivers ofdistribution and risks to soil fauna of seven metals (Cd, Cu,Cr, Ni, Pb, V and Zn) in British soils,” Environmental Pollution,vol. 153, no. 2, pp. 273–283, 2008.

[24] R. Swennen, I. VanKeer, and W. DeVos, “Heavy metal con-tamination in overbank sediments of the Geul River (eastBelgium): its relation to former Pb-Zn mining activities,” En-vironmental Geology, vol. 24, no. 1, pp. 12–21, 1994.

[25] H. Leenaers, “The transport of heavy metals during floodevents in the polluted River Geul (The Netherlands),” Hydro-logical Processes, vol. 3, no. 4, pp. 325–338, 1989.

[26] B. Bølviken, J. Bogen, W. De Vos et al., “Final report of theWorking Group on Regional Geochemical Mapping 1986-1993. Forum of European Geological Surveys (FORGES),”Geological Survey of Norway (NGU) Open File Report 93.092,1993.

[27] B. Bolviken, J. Bogen, A. Demetriades et al., “Regional geo-chemical mapping of Western Europe towards the year 2000,”Journal of Geochemical Exploration, vol. 56, no. 2, pp. 141–166,1996.

[28] R. Salminen, M. J. Batista, M. Bidovec, and A. Demetriades,Geochemical Atlas of Europe. Part 1: Background Information,Methodology and Maps, EuroGeosurveys & Foregs, Espoo,Finland, 2005.

[29] R. Salminen, T. Tarvainen, A. Demetriades et al., FOREGSgeochemical mapping field manual, Geological Survey ofFinland, Guide 47, 1998.

[30] D. W. Nelson and L. E. Sommers, “Total carbon, organic car-bon and organic matter,” in Methods of Soil Analysis. Part 2:Chemical and Microbiological Properties, pp. 516–593, Amer-ican Society of Agronomy, Madison, Wis, USA, 2nd edition,1982.

[31] P. Quevauviller, “Operationally defined extraction proceduresfor soil and sediment analysis I. Standardization,” TrAC, vol.17, no. 5, pp. 289–298, 1998.

[32] H. Sandstrom, S. Reeder, A. Bartha et al., “Sample preparationand analysis,” in Geochemical Atlas of Europe. Part 1—Back-ground Information, Methodology, and Maps, R. Salminen etal., Ed., Geological Survey of Finland, Espoo, Finland, 1998.

[33] L. De Temmerman, L. Vanongeval, W. Boon, M. Hoenig, andM. Geypens, “Heavy metal content of arable soils in NorthernBelgium,” Water, Air, and Soil Pollution, vol. 148, no. 1–4, pp.61–76, 2003.

[34] R. Webster, “Regression and functional relations,” EuropeanJournal of Soil Science, vol. 48, no. 3, pp. 557–566, 1997.

[35] R. Webster, “Statistics to support soil research and their pres-entation,” European Journal of Soil Science, vol. 52, no. 2, pp.331–340, 2001.

[36] Z. Peh and S. Miko, “Geochemical comparison of stream andoverbank sediments: a case study from the zumberak region,Croatia,” Geologia Croatica, vol. 54, no. 1, pp. 119–130, 2001.

[37] G. M. Ivanov and V. K. Kashin, “Vanadium in the landscapesof western Transbaikalia,” Geochemistry International, vol. 48,no. 3, pp. 295–299, 2010.

[38] A. J. Wigginton and D. J. Price, “Analysis of Metals in Water,Stream Sediments, and Floodplain Soils Collected May 23-25, 2006 from the Bayou Creek System,” DRAFT REPORT,February 2007.

[39] W. M. Mayes, A. P. Jarvis, I. T. Burke et al., “Dispersal andattenuation of trace contaminants downstream of the ajkabauxite residue (Red Mud) depository failure, Hungary,” Envi-ronmental Science and Technology, vol. 45, no. 12, pp. 5147–5155, 2011.


[40] L. R. Lado, T. Hengl, and H. I. Reuter, “Heavy metalsin European soils: a geostatistical analysis of the FOREGSGeochemical database,” Geoderma, vol. 148, no. 2, pp. 189–199, 2008.

[41] L. Horckmans, R. Swennen, J. Deckers, and R. Maquil, “Localbackground concentrations of trace elements in soils: a casestudy in the Grand Duchy of Luxembourg,” Catena, vol. 59,no. 3, pp. 279–304, 2005.

[42] P. Tume, J. Bech, L. Tume et al., “Concentrations anddistributions of Ba, Cr, Sr, V, Al, and Fe in Torrelles soil profiles(Catalonia, Spain),” Journal of Geochemical Exploration, vol.96, no. 2-3, pp. 94–105, 2008.

[43] D. J. Huisman, F. J. H. Vermeulen, J. Baker, A. Veldkamp, S.B. Kroonenberg, and G. T. Klaver, “A geological interpretationof heavy metal concentrations in soils and sediments in thesouthern Netherlands,” Journal of Geochemical Exploration,vol. 59, no. 3, pp. 163–174, 1997.

[44] M. Aide, “Geochemical assessment of iron and vanadium rela-tionships in oxic soil environments,” Soil and Sediment Con-tamination, vol. 14, no. 5, pp. 403–416, 2005.

[45] S. R. Oliva and A. J. F. Espinosa, “Monitoring of heavy metalsin topsoils, atmospheric particles and plant leaves to identifypossible contamination sources,” Microchemical Journal, vol.86, no. 1, pp. 131–139, 2007.

[46] T. Sterckeman, F. Douay, D. Baize, H. Fourrier, N. Proix, andC. Schvartz, “Trace elements in soils developed in sedimentarymaterials from Northern France,” Geoderma, vol. 136, no. 3-4,pp. 912–929, 2006.

[47] A. Naeem, P. Westerhoff, and S. Mustafa, “Vanadium removalby metal (hydr)oxide adsorbents,” Water Research, vol. 41, no.7, pp. 1596–1602, 2007.

[48] B. Wehrli and W. Stumm, “Vanadyl in natural waters: ad-sorption and hydrolysis promote oxygenation,” Geochimica etCosmochimica Acta, vol. 53, no. 1, pp. 69–77, 1989.

[49] VLAREBO, “Vlaams reglement betreffende de bodemsaner-ing. Decree on Soil Remediation and Soil protection,” ratifiedby the Flemish Government on October 2006.

[50] C. Carlon, M. Dalla Valle, and A. Marcomini, “Regressionmodels to predict water-soil heavy metals partition coefficientsin risk assessment studies,” Environmental Pollution, vol. 127,no. 1, pp. 109–115, 2004.

[51] D. G. Schulze, “Separation and concentration of iron-con-taining phases,” in Iron in Soils and Clay Minerals, J. W. Stucki,B. A. Goodman, and U. Schwertmann, Eds., pp. 63–81, Reidel,Dordrecht, The Netherlands, 1988.

[52] S. Covelli and G. Fontolan, “Application of a normalizationprocedure in determining regional geochemical baselines,” En-vironmental Geology, vol. 30, no. 1-2, pp. 34–45, 1997.

[53] K. C. Schiff and S. B. Weisberg, “Iron as a reference element fordetermining trace metal enrichment in Southern Californiacoastal shelf sediments,” Marine Environmental Research, vol.48, no. 2, pp. 161–176, 1999.

[54] V. J. G. Houba, T. M. Lexmond, I. Novozamsky, and J. J. VanDer Lee, “State of the art and future developments in soilanalysis for bioavailability assessment,” Science of the TotalEnvironment, vol. 178, pp. 21–28, 1996.

[55] M. McBride, S. Sauve, and W. Hendershot, “Solubility controlof Cu, Zn, Cd and Pb in contaminated soils,” European Journalof Soil Science, vol. 48, no. 2, pp. 337–346, 1997.

[56] S. Sauve, W. Hendershot, and H. E. Allen, “Solid-solutionpartitioning of metals in contaminated soils: dependence onpH, total metal burden, and organic matter,” EnvironmentalScience and Technology, vol. 34, no. 7, pp. 1125–1131, 2000.

[57] K. L. Mandiwana, N. Panichev, and R. Molatlhegi, “Theleaching of V(V) with PO 3−

4 in the speciation analysis of soil,”Analytica Chimica Acta, vol. 545, no. 2, pp. 239–243, 2005.

[58] W. P. Miller, D. C. Martens, L. W. Zelazny, and E. T. Kornegay,“Forms of solid phase copper in copper-enriched swine ma-nure,” Journal of Environmental Quality, vol. 15, no. 1, pp. 69–72, 1986.

[59] O. K. Borggaard, “Selective extraction of amorphous ironoxides by EDTA from a Danish sandy loam,” Journal of SoilScience, vol. 30, pp. 727–734, 1979.

[60] B. Nowack and L. Sigg, “Dissolution of Fe(III)(hydr) oxidesby metal-EDTA complexes,” Geochimica et Cosmochimica Acta,vol. 61, no. 5, pp. 951–963, 1997.

[61] H. E. Gabler, K. Gluh, A. Bahr, and J. Utermann, “Quantifi-cation of vanadium adsorption by German soils,” Journal ofGeochemical Exploration, vol. 103, no. 1, pp. 37–44, 2009.

applied and environmental soil science impact of human activities

Documents