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Transport of complexed cyanide in soil
ELSEVIER
Geoderma67 ( 1995) 73-85
Transport of complexed cyanide in soil J.C.L. Meeussen *, W.H. van Riemsdijk, S.E.A.T.M. van der Zee Department of Soil Science and Plant Nutrition, Wageningen Agricultural University, P.O. Box 800.5. 6700 EC Wageningen, The Netherlands
Received 15 March 1994; accepted after revision 2 September 1994
Abstract Contamination and Fe(CN):-1,
of soil with cyanide, generally in the form of iron cyanide complexes [Fe( CN)iis commonly found at several types of industrial sites. The risks for human health
or the environment posed by such sites are largely determined by the chemical behaviour and transport of complexed cyanide in soil. In acidic soil this behaviour is probably dominated by equilibrium with Prussian blue, Fe,(Fe(CN),),(s), which is sparingly soluble under acidic conditions and limits dissolved cyanide concentrations and mobility. However, atpH levels higher than about 7 the solubility of this precipitate is extremely high, which allows cyanide to be very mobile under such conditions. Nevertheless, according to field observations, Prussian blue appears to persist for decades in alkaline soils. In this paper multicomponent transport calculations, including equilibrium with Prussian blue, iron hydroxide and relevant redox reactions, are compared with the situation as observed in two contaminated soil profiles. In case of an acidic soil, there appears to be good agreement between the calculations and the field situation. Here the mobility of cyanide is mainly regulated by the pH and redox potential and is relatively low. In case of an alkaline soil, precipitation of Prussian blue will not take place and cyanide will be very mobile once leached from the initial waste material. In this case the dissolution rate of Prussian blue is the process which determines cyanide concentrations in the groundwater. The calculations show that this dissolution rate is determined by the acid neutralizing capacity of the solution infiltrating in the waste material. The buffer capacity depends on the soil composition and can easily be so low that complete dissolution of Prussian blue may take decades. This very likely explains the persistence of this material in alkaline soils.
1. Introduction The use of cyanide in several kinds of industrial processes has regularly caused local contamination of the soil with this substance. It is commonly found as a soil contaminant * Corresponding author. Present address: The Macaulay Land Use Research Institute, Craigiebuckler, Aberdeen AB9 2QJ, Scotland, UK. 0016-7061/95/$9.50
0 1995 Elsevier Science B.V. All rights reserved
SsDIOOI6-7061(94)00061-1
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at the sites of electroplating industries, metal ore processing industries and of coal gasification plants. Especially at the sites of (former) coal gasification plants, of which there are over 200 in the Netherlands, cyanide is often present in large amounts. These coal gasification plants have produced large amounts of cyanide containing waste material, which originated from the gas purification process. This waste material contained high concentrations of cyanide in the form of Prussian blue. Because this material was more or less worthless it was often stored or disposed of at the surface of the sites close to the gasification plants. In numerous occasions the presence of the waste material has caused contamination of the soil with cyanide by leaching of dissolved iron cyanide, or by contamination of the soil with the solid waste material itself. Because cyanide is a potentially highly toxic substance its presence in the soil may pose serious risks for human health and for the environment when it enters the ecological cycle by leaching to the groundwater. The risks for such hazards strongly depend on the bioavailability and mobility of cyanide in the soil. Both properties are strongly influenced by the chemical forms in which cyanide is present, and by the way these forms are bound to the soil solid phase. When cyanide was disposed of in the environment, in most cases it was discharged in the form of iron cyanide complexes [Fe(CN)iand Fe(CN)z- 1. These complexes are not very toxic (Fuller, 1984) but when exposed to daylight they readily decompose to highly toxic and volatile free cyanide [ HCN( aq) ] : Fe(CN)il-+6H,O+3H+%Fe(OH),(s)+6HCN(aq)
(1)
Free cyanide is not only toxic and volatile, but due to its uncharged nature also very mobile in soils ( Alesii and Fuller, 1976). Although free cyanide is susceptible to biodegradation in low concentrations, it is clearly the more hazardous species when compared to the complexed forms. According to thermodynamic equilibrium calculations, iron cyanide complexes tend to decompose to free cyanide in soils. Fortunately, this decomposition proceeds extremely slowly in darkness and thus in soils. In fact the decomposition rate of iron cyanide in soils is so low that it makes iron cyanide kinetically stable, and allows it to remain the predominant cyanide species in soils (Meeussen et al., 1992a). In case soil contamination has occurred with cyanide in the form of iron cyanide complexes, the behaviour of these complexes determines the overall behaviour of cyanide in the soil. This behaviour is governed by the interaction between dissolved iron cyanide complexes and the soil solid phase. In case of complexed cyanide the principal interaction with the soil solid phase is by precipitation and dissolution of Prussian blue, a mineral consisting of iron cyanide complexes and iron [ Fe,( Fe( CN),) S( s) ] (Meeussen et al., 1992b). The extra iron necessary for the formation of this mineral is generally abundantly present in soils, e.g. in the form of iron (hydr)oxides. Dissolution of Prussian blue can be represented by the following reaction: Fe4(Fe(CN),)-,(s)
%iFe(CN)i-
+4Fe,+
+3e-
(2)
Under aerobic conditions, the evolving Fe’ ?? will precipitate as an iron (hydr)oxide, e.g. Fe(OH),( s). This results in the following overall dissolution reaction for Prussian blue: Fe,(Fe(CN),),(s)
+ 12H,Or;i4Fe(OH),(s)
+3Fe(CN)i-
+3e-+
12H’
(3)
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According to this reaction, dissolution of Prussian blue consumes hydroxyl ions and produces electrons and therefore has an acidifying and reducing effect on the local environment. When Prussian blue dissolves under anaerobic conditions and little or no electron consuming agents are present, the evolving iron will partly remain in the reduced (Fe*+) state and more reduced states of iron hydroxide will form, e.g. Fe, (OH)*(s) (Schwab and Lindsay, 1983). This leads to the following overall dissolution reaction:
(4) The way that protons and electrons are involved in these dissolution reactions indicates that the solubility of Prussian blue depends on soil pH as well as on redox potential. In general a low pH and a low redox potential result in a low solubility of Prussian blue, and thus in low cyanide concentrations in the soil water phase and in a limited mobility of complexed cyanide. Additionally, precipitation or dissolution of Prussian blue may influence the pH and redox potential of the soil. The significance of this effect depends on the amount of Prussian blue present and on the transport behaviour of acid and base in the soil. The transport of acid and base in soils is governed by the buffering characteristics of the soil and the soil solution (Scheidegger et al., 1994). From the overall reaction equations it appears that four hydroxyl ions are needed to dissolve one mole of Fe(CN), from Prussian blue in case of reaction 3, and about 3.5 in case of reaction 4. This makes the solubility of Prussian blue extremely pH-dependent. According to simple equilibrium calculations, this precipitate would be completely soluble and thus mobile at the pH levels found in calcareous soils. Nevertheless, solid Prussian blue often appears to persist in such soils for at least decades after its disposal in waste material. In contrast with the situation in the acidic soils, where equilibrium calculations give reasonable predictions of occurring concentrations, the situations found in alkaline soils therefore cannot be explained by simply assuming a chemical equilibrium situation. The main aim of this work is to evaluate whether existing knowledge of the behaviour of complexed cyanide in soil and in the dissolution and precipitation of Prussian blue is adequate for modeling the transport of iron cyanide in contaminated soils, and to elucidate the cause for the persistence of Prussian blue in alkaline soil. This evaluation is made by comparing transport and accumulation of iron cyanide as observed at two contaminated sites to calculations with a multicomponent transport model in which the relevant chemical equilibrium reactions were incorporated. A further objective was to use this transport model for the identification of soil parameters that have a large impact on mobility of cyanide in soils. 2. Methods and materials 2. I. Field data
Field data on total cyanide content of the soil, pH, redox potential and dissolved cyanide concentrations were obtained by sampling two sites of former coal gasification plants (A
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and B). The collection of these field data has been described previously (Meeussen et al., 1994). The two sites were selected primarily because of their difference in soil type. Site A consists of an acidic sandy soil while at site B a calcareous clay soil is present. Both soil types are commonly prevalent in the Netherlands. The pH of the soil at the acidic site (A) was about 4.1, while the pH of the calcareous soil at site B was about 7.5. Because of the huge effect of pH on the dissolution and precipitation of Prussian blue, there is a substantial difference in behaviour of cyanide between these soils. At the surface of site A, the acidic site, no waste material in its original form remained at the time of the investigations. The top soil had apparently been replaced by clean soil material. At many contaminated sites the conditions are presently similar, especially when such sites are used for non-industrial purposes, e.g. for a residential area, which was the case here. The cyanide contamination of the soil present below the top soil layer was most likely caused by leaching of cyanide from waste material that previously covered the surface. The present soil consists of an aerobic layer that contained little cyanide and an anaerobic, water saturated, layer where cyanide had apparently accumulated. The accumulation was easily detectable due to the blue colour of this layer. On site B, the more alkaline site, the conditions were quite different. Here the original gasworks building was still present, although now used for other purposes. The surface of this site most likely has not changed much since gas production was discontinued here about thirty years ago. At least some of the original waste material was still present here, which was illustrated by clearly visible blue spots at the soil surface and by the high cyanide concentrations found in the top soil layer. In contrast with the acidic site, the blue colour was not observed in the soil layers below the surface. In the deeper layers there was virtually no cyanide present in the soil solid phase. However, high concentrations were found here in the soil water. 2.2. Calculations Calculations of transport and accumulation of iron cyanide were done using ECOSAT, a multicomponent transport computer model (Keizer et al., 1993). This model consists of a chemical module which calculates the speciation of chemical components and their distribution over different species and phases, and a transport module which calculates convective and dispersive transport of dissolved species. Both parts of the computer program are separated as much as possible. This allows the use of more sophisticated chemical or transport modules, and makes replacement of either of the modules by alternative ones relatively easy. This alternate calculation of chemical equilibrium and mass transport can cause extra numerical dispersion. However, in the ECOSAT program this dispersion is diminished by iteration between both modules until convergence is reached. The chemical reactions included in the model calculations are listed in Table 1. The most important assumption in modelling the transport of cyanide in this way is that interaction of dissolved iron cyanide with the soil solid phase only takes place by precipitation of Prussian blue (Table 1, reaction 3). Other possible interactions with the soil solid phase, for example adsorption of iron cyanide, were not considered. Another assumption was that in the acidic soil the buffer capacity for protons and electrons was completely supplied by dissolution and precipitation of iron (hydr) oxide minerals and of Prussian blue.
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67(1995)
73-85
71
Table 1 Equilibrium reactions and constants used in the transport calculations Reaction
log K”
Reference
3.54 4.63 - 84.5
Lindsay ( 1979) Lindsay ( 1979) Meeussen et al. ( 1992b)
13.0 6.3 - 2.2 -5.7 - 3.0 - 19.0 - 13.1 -21.6
Lindsay Lindsay Lindsay Lindsay Lindsay Lindsay Lindsay Lindsay
Solid phases
(1) Fe(OH),(s) +3H+ %Fe3++3H20 (2) Fe,(OH),(s)+8H+ S3Fe3++e-+8Hz (3) Fe,(Fe(CN),)3(s) %4Fe’* +3Fe(CN)zIron hydroxo complexes (4) Fe’+ +e-%Fe’+
(5) Fe3++H,0+e-%Fe(OH)+ +H+ (6) Fe3++HZOSFe(OH)*++Hf (7) Fe3++2H,0%Fe(OH):+2H’ (8) Fe3++2H,0+e-%Fe(OH)i+2Ht (9) Fe’++3H,O+e-%Fe(OH), +3H+ (10) Fe’+ +3H,O%Fe(OH)t+3H+ (11) Fe3++4H,0%Fe(OH),+4Ht Fe(CN), complexes (12) Fe(CN)z-+e-%Fe(CN)z-
(13) Fe(CN)i- +e-+H+SHFe(CN)i(14) Fe(CN)i-+e-+2H+ QH,Fe(CN)z-
+3e-
6.0 10.4 12.8
( 1979) (1979) ( 1979) ( 1979) ( 1979) ( 1979) ( 1979) ( 1979)
Beck (1987) Beck (1987) Beck (1987)
Carbonate equilibria
(15) (16) (17) (18)
CaC03(s) %Ca*+ +CO:CO:-+H+ %HCO, HCO; +H+ SH,CO; CO,(g) +HzO%H2CO;
-8.4 10.3 6.4 - 1.5
Lindsay Lindsay Lindsay Lindsay
( 1979) (1979) ( 1979) ( 1979)
In reality, of course, there are numerous other processes that may supply this kind of buffer capacity. The possible effects of this simplification are mentioned in the results and discussion section. In order to perform the calculations, the initial column composition with respect to total amounts of substances and initial pH of the solution was given as input. During the calculations the pH and redox potential were calculated, so these parameters were not fixed or manually adjusted. In case of the calculations with an acidic soil, the initial situation in the modelled column consisted of three separate zones. Although no longer present in reality, on top we located a layer of 0.1 m of waste material consisting of 0.01 M l- ’solid Prussian blue and 0.1 M 1- ’Fe( OH),( s), serving as source for cyanide leaching. Between this layer and the groundwater table we located an oxidized zone consisting of 0.1 M 1-i Fe(OH),( s) only, and below this layer a more reduced zone, where a mixture of 0.05 M l- ’ Fe( OH), (s) and 0.05 M I- ’Fe,( OH) s( s) was present. The amount of protons initially present in the soil was chosen to establish a pH of 4. The infiltrating solution consisted of water with a pH of 4 and did not contain cyanide. In the alkaline soil case, the initial situation in the modelled column was similar. Again three different zones were distinguished. A layer of 0.1 m of waste material, containing 0.01 mole Prussian blue l- 1 and 0.1 M I-’ Fe(OH),(s), was placed on top. Below this layer was located a homogeneous layer containing 0.1 M 1-l Fe( OH),( s). The deepest
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J.C.L. Meeussen et al. / Geodena
67 (1995) 73%RS
layer consisted of a reduced zone where, like in the acidic soil, 0.05 M 1-l Fe(OH),( s) and 0.05 M 1-l Fe,(OH),( s) was present. In this case, the pH of the infiltrating solution was adjusted to 7.0, and the initial pH of the soil was adjusted to 7.2. In order to simulate the alkaline nature of the soil in this column, 0.1 M l- ’CaCO, in equilibrium with atmospheric CO, was added to the two layers below the contaminated layer. Without such a proton buffer the entire profile was likely to acidify within only a few pore volumes. As discussed below, calcite was not added directly to the contaminated layer because it is not stable in contact with Prussian blue. The transport calculations considered a soil column having a volumetric water content of 0.3 (water saturated), a water flux density of 0.35 m yr- ’ and for a period of 30 years of leaching.
3. Results and discussion 3. I. Acidic soil The cyanide content, pH and redox potential (or pe = 16.95&) as measured in the acidic contaminated soil are shown as a function of depth in Fig. 1, as are also the calculated values for these parameters after simulation of a thirty-year leaching period. The modelled situation shows a heavily contaminated top layer. If we first consider the pH profile, it Soil A measured ---- calculated
depth (cm) 0
-50
,_______..._...----.-.
I
j
$ -------
waste material
-100
-250
\
j
*
anaerobic zone
table 3
CN(t) soil mglkg Fig. 1. Profiles of total cyanide concentrations, calculated with the transport model, for soil A.
4
5
6
7
6
10
8
PH pH and redox potential as measured
12
pe in a field situation and as
J. C. L. Meeussen et al. / Geoderma 67 (1995) 73-85
79
appears that the modelled pH values are very close to the values observed in the field. The pH was initialized at 4 at the start of the calculations, but this is no guarantee for the pH to remain around this value during the calculations. The calculations show a slight pH increase in the aerobic zone ( f0.2 units) compared to the initial pH of 4. This increase occurs during the first few pore volumes of leaching and is the result of the precipitation of a small amount of Prussian blue. Hereafter, in this zone the pH remains constant and no further precipitation of Prussian blue occurs. At the top of the anaerobic layer a more profound increase of the pH can be observed ( f 0.5 units). A comparable increase is observed in the field situation. Another relevant parameter is the redox potential of the soil solution. As also observed in the case of pH, it appears that the modelled results agree quite well with the field measured values, in the aerobic as well as in the anaerobic zone. In general, correct determination of redox potentials in soil solutions is difficult, and observed values are usually not very reliable. However, determination of the redox potential is relatively reliable in solutions containing Fe(CN), ions, due to the fast redox equilibrium between Fe(CN)i- and Fe(CN)z-. The calculated pH and redox values are sufficiently close to the real values to allow prediction of the location of dissolution of Prussian blue in the soil, although the pH and redox buffering mechanisms in the real soil are without doubt much more complex than the simple precipitation reactions of the different minerals as implemented in the model. In the field situation, iron cyanide appears not to accumulate significantly in the aerobic zone. Hence, only very small amounts of iron cyanide were found here in the soil solid phase. The same outcome was observed in the modelled results. During the calculations the initial conditions in the column favoured precipitation of Prussian blue in the aerobic zone and initially some Prussian blue precipitated there. However, this resulted in an increased pH which prevented further precipitation. In this way, only a small part of the leached iron cyanide precipitated and became immobile while the rest of the iron cyanide remained available for transport and thus was mobile in this layer. In the calculations, the exact amount of iron cyanide that precipitates in this layer is determined by the initial pH of the soil solution. In the field this amount is strongly influenced by the soil’s proton buffer capacity. If this buffer capacity is large enough, precipitation of Prussian blue does not affect pH, and complete precipitation may occur. In the anaerobic zone, especially in the upper part of it, accumulation of iron cyanide is observed both in the field situation and in the calculations. In this layer the solubility of Prussian blue is lower than in the aerobic zone because of the lower redox potential. The Prussian blue that precipitates here tends to increase the redox potential and the pH. In the model, this effect is neutralized by the transformation of Fe,(OH)8 to Fe( 0H)3, which buffers the redox potential. Therefore, the solubility of Prussian blue remains low and accumulation is possible in this layer. When the conversion of Fe,(OH), to Fe(OH), is complete, no further precipitation of Prussian blue occurs. It is possible that in the field situation the same processes buffer the pH and redox conditions, but it is very likely that here organic matter will also act as a reducing agent, making continued precipitation of iron cyanide possible. The layer where accumulation of cyanide occurs in the field is more diffuse than calculated with the model. This is probably mainly caused by fluctuations in the groundwater table in the field, which causes the redox profile here to be more diffuse than in the model calculations.
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J.C.L. Meeussen et al. /Geodertna 67 (1995) 73-B
Soil B depth (cm) ---
measured calculated
I:
._-_-__.-__._.__-._.-
-200
groundLater table I I 1
-250
1
500
CN(t) soil mglkg
1000
3
4
5
PH
6
7
6
8
10
12
pe
Fig. 2. Profiles of total cyanide concentrations, pH and redox potential as measured in the alkaline field situation and as calculated with the transport model, for soil B.
3.2. Alkaline soil The observed situation in the alkaline field soil differs in several aspects from that in the acidic soil. First, the highest concentration of soil cyanide was here present in the top soil layer, which most likely contained some original waste material. The blue colour of this layer strongly suggested the presence of Prussian blue. In the deeper layers, where much less cyanide was present, this colour was absent. Most of the cyanide in these layers was found in the dissolved phase (Meeussen et al., 1994). The concentration of iron cyanide measured in the soil water of the saturated zone at this site was more than ten times higher thaninthesoilwateroftheacidicsite[1.1X10~5MFe(CN),comparedto9.3X10~7M Fe( CN),] . When we look at the pH profile of the soil (Fig. 2) it appears that the pH at this site is significantly higher than at the acidic site. There is also a more pronounced difference in pH between the layers, with pH of the top soil about 5.6, while in the deeper layers it is about 7.2. 3.3. Modelling results Let us first consider the results of the calculations with the alkaline soil, which had an initial pH of the column of 7.2, and the addition of calcite to the subsurface soil layers to buffer the pH there. Immediately after the calculations were started, the pH in the top layer decreased to about 5.2, due to the dissolution of Prussian blue. A very similar pH increase with depth was observed in the field, although pH in the upper soil layer was slightly higher here.
J. C. L Meeussen et al. / Geoderma 67 (I 995) 73-%5
81
The pH of the leachate was buffered to about 7.3 by dissolving calcite in the deeper layers. The dissolved iron cyanide that leached from the contaminated layer did not reprecipitate at the alkaline pH levels in the deeper layers and remained mobile. Dissolution of Prussian blue in alkaline soils is clearly a one direction process. This is largely in agreement with the field observations where precipitated iron cyanide was also present only in the contaminated top layer. In these calculations, the concentrations of dissolved iron cyanide which can occur in the ground water depend mainly on the dissolution rate of Prussian blue in the waste material (upper layer). The acidification of this layer limits the solubility of iron cyanide in the calculations to about 5 X 10e6 M Fe(CN)6 in case of leaching with pure water, which is about the same concentration as observed and calculated in the case of the acidic soil. However, the iron cyanide concentrations found in the alkaline soil water are well above this concentration. A possible explanation for this phenomenon follows. One of the main reasons for the higher pH at the alkaline site is the occurrence of calcite. Dissolution of this mineral buffers the pH at an alkaline level. Calcite can also react with Prussian blue according to the following reaction: Fe,(Fe(CN),),(s) *4Fe(OH),(s)
+6CaCO,(s) +3Fe(CN)z-
+ 12H,O +6H2C03+6Ca2+
(5)
According to chemical equilibrium calculations, this acid base reaction proceeds completely, which implies that Prussian blue and calcite are not chemically stable when mixed. One of the two minerals will dissolve entirely, depending on which of them is more abundant. In that case, the resulting concentration of dissolved iron cyanide is equal to half the concentration of dissolved calcium. An alkaline pH will develop when there is an excess of calcite, while an acidic pH will evolve when there is more Prussian blue present. According to the acidic pH of the top soil layer, calcite is most likely absent here, while according to the measurements and the soil colour Prussian blue is absent in the deeper layers. When solid Prussian blue is spatially separated from solid calcite, the reaction between both minerals, and thus dissolution of both, can still proceed via reaction with mobile dissolved ions. When calcite reacts with dissolved carbon dioxide it dissolves according to the following reaction: CaCO, ( s) + H.&O, * 2HCO; + Ca*+
(6)
The evolving dissolved ions are able to migrate to solid Prussian blue, where the following reaction takes place: Fe,(Fe(CN)&(s)
+ 12HCO; + 12H,O%4Fe(OH),(s)
+3Fe(CN)z-+6H,CO, (7)
The combination of reaction 6 (6 times) with reaction 7 is equivalent to reaction 5. The maximum rate at which this reaction proceeds is, apart from pure chemical dissolution kinetics, limited by the rate at which especially the HCO; ion is transported from one mineral to the other. In case of short distances between these solid materials, this transport rate will be governed by diffusion. An expression for the flux density in case of diffusion is given by Fick’s law for molecular diffusion, with modifications for porosity and pore tortuosity:
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67 (1995)
J= 8.D.(Sc/Sx)
73-85
(8)
where J is the flux density (in mol m-2 s- ’), 8 is the porosity, D is the diffusion coefficient in soil, which is derived from the molecular diffusion coefficient in free solution after correction for the tortuosity, which has a value of about 1.25 -9 m2 s- ‘, c is the concentration (in mol m- “), and x the distance (in m) (Bolt, 1982). The factor that determines the transport rate in this case is the gradient of ion concentration. Because the concentration differences are relatively constant due to buffering by the dissolving minerals, the diffusion rate will be predominantly determined by the distance between the solid phases. In case of short distances between both solids, e.g. when both minerals are present in very small particles and are intensively mixed, these diffusion fluxes are high, and allow fast dissolution of both minerals. However, under such conditions dissolution rates might be limited by pure chemical dissolution kinetics. Apart from diffusion, the transport of dissolved ions can also result from convection. In humid climates this is the dominant transport process. In case of convective transport the flux density equation equals: J= 8.11.~
(9a)
where J is the flux density (in mol rnp2 s- ’), 19is the porosity, c is the concentration (in mol m- 3), and u is the the pore water velocity (in m s - 1) . However, the effective convective flux of dissolved ions to (or from) a solid particle is determined by the difference in concentration between the solution before and after contact with the particle. In case of instantaneous equilibrium this is the difference between the concentration in the bulk solution and the concentration in equilibrium with the solid phase. If we consider this concentration difference to be 6c, this leads to: J= e.u.8~
(9b)
By combining Eqs. (8) and (9a) it is possible to estimate the distance x, at which convective transport becomes predominant over diffusive transport. From the combination of both equations follows the expression for the condition where the diffusive flux equals the effective convective flux: e.D.(6c/Sx)
=e.Lj,sc
( 10)
If we consider the distance to be x this can be reduced to:
(11) The net infiltration rate in the Netherlands on average equals 0.35 m yr- ‘, which results for water saturated conditions in a pore water velocity u of 3.6 X lo-’ m s- ’) . The unknown distance x is found by using Eq. ( 11) which results in 0.035 m. This implies that convective transport of ions dominates when solid calcite is separated more than about 3.5 centimetres from solid Prussian blue. This calculation is based on an averaged water velocity. However, convective water fluxes during wetter periods in the field can be orders of magnitudes higher, which makes convective transport already dominant at much shorter distances (Bolt, 1982).
J.C.L. Meeussen etal. / Geodemza 67 (I 995) 7345
83
0
-2 -
-3 -
-4 -
-5 -
-6
I -7
I -6
I -5
I -4
log alkalinity
I -3
I -2
I -1
I 0
(mol/l)
Fig. 3. Calculated concentration of dissolved cyanide in leachate from Prussian blue containing waste material as function of the alkalinity of the infiltrating solution.
Because original waste material was not well mixed with the soil, convective transport of alkalinity will in most cases be the dominating process regulating the dissolution of Prussian blue. In that case the concentration of iron cyanide that results from dissolution depends on the capacity of the infiltrating solution to neutralize acid. The acid neutralizing capacity is determined by the alkalinity of the solution, which equals [ OH - ] + [ HCO; ] + 2 [CO: - 1. If we assume that an infiltrating solution equilibrates with Prussian blue while passing a contaminated layer, the resulting iron cyanide concentrations can easily be calculated with the ECOSAT model (Fig. 3). At high alkalinities, the concentration of Fe( CN)z- in the leachate is equal to four times the alkalinity because four moles of alkalinity are needed to dissolve one mole of Fe(CN)z-. At low alkalinities, the minimum concentration of dissolved iron cyanide is limited by the solubility of Prussian blue in pure water (which has an alkalinity of zero). The concentrations in Fig. 3 represent maximum concentrations, because in the field heterogeneity can cause incomplete contact between leaching water and Prussian blue. Nevertheless, with these assumptions a maximum value for iron cyanide concentrations in the leachate of cyanide-containing waste material can be derived. In calcareous soils, the alkalinity of the soil water, is strongly influenced by equilibrium between calcite and atmospheric carbon dioxide, and ranges from 1O-4 to 10e3 M (Lindsay, 1979). If this soil water leaches through the Prussian blue containing layer, iron cyanide concentrations of about 4 X lop5 M to 2 X 10e4 M will result, which are higher than observed in the field. If there is no equilibrium with calcite in the upper layer, the alkalinity of soil solution in the top soil layer is probably significantly lower. If we assume an alkalinity of low5 M, for which a CO2 pressure of 3 mbar is needed at the pH of the top soil, this results in concentrations of dissolved cyanide which are in the range of the concentrations as observed in the
84
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67 (1995) 73-85
field. Differences between observations and calculations can very well be caused by mixing of contaminated groundwater with clean groundwater or because either not all of the infiltrating water is in equilibrium with calcite or it equilibrates with Prussian blue while leaching. Each of these factors may lower the concentration of iron cyanide in leachates. Nevertheless, these assumptions can be used to estimate the time necessary to dissolve all Prussian blue in this soil. When we consider the present situation at site B, it is possible to estimate the time it will take before dissolution is complete if dissolution rates remain constant in time. At site B, a layer of about 10 cm thickness with a total cyanide content of 930 mg CN kg- ’is present. This is equal to 6 X 10M3 M Fe( CN), kg- ’which makes about 1 X 10-2MFe(CN)61-’ at anestimatedbulkdensityof 1500 kg m-‘. Underthe hydrologic conditions in the Netherlands, with a net water flux of 0.35 m yr~ ’ and a concentration in theleachateof1.1X10~5,itwouldtake(1X10~2)/(1.1X10~5X3)=300year.Thisis clearly long enough to justify the survival of Prussian blue in an alkaline soil for over thirty years. At the calculated maximum leaching concentration of 2 X 10e4 M l- ’ complete dissolution of the cyanide would take about 15 years.
4. Conclusions
-
-
-
Observed accumulation of iron cyanide in the field qualitatively agrees with transport calculations based on the assumption that interaction of dissolved cyanide with the soil solid phase is dominated by precipitation of Prussian blue. In soils with pH levels below about 5, concentrations of dissolved iron cyanide are governed by equilibrium with Prussian blue. Prussian blue is not stable and will eventually dissolve in alkaline soils. Here, dissolution rates of Prussian blue and the resulting concentrations of dissolved iron cyanide will be largely governed by the alkalinity of the infiltrating solution. To leach iron cyanide from a contaminated soil for remediation purposes, a solution with a high alkalinity is needed. Under field conditions leaching rates in alkaline soils can easily be so low that complete leaching of iron cyanide from contaminated soils can take decades.
References Al&i, B.A. and Fuller, W.H., 1976. The mobility of three cyanide forms in soils. EPA-600/9-76-015, Residue Management and Land Disposal, Proc. Hazard. Waste Res. Symp., pp. 213-223. Beck, M.T., 1987. Critical evaluation of stability constants of metal complexes in solution. Critical survey of cyano complexes. Pure Appl. Chem., 59: 1703-1720. Bolt, G.H., 1982. Soil Chemistry, B. Physic0 Chemical Models. Elsevier, Amsterdam. Fuller, W.H., 1984. Cyanide in the environment with particular attention to Soil. In: D. Van Zyl (Editor), Cyanide and the Environment, I. Geochemical Engineering Program. Colorado State Univ., Fort Collins, pp. 1946. Keizer,M.G., De Wit, J.C.M., Meeussen, J.C.L.,Bosma, W.J.P.,Nederlof, M.M., Venema,P., Meeussen, V.C.S., Van Riemsdijk, W.H and Van der Zee, S.E.A.T.M., 1993. ECOSAT, A computer program for the calculation
J.C.L. Meeussen et al. /Geodenna
67 (1995) 73-85
85
of chemical speciation and transport in soil-water systems. Dep. of Soil Science and Plant Nutrition, Wageningen Agricultural University, Wageningen. Lindsay, W.L., 1979. Chemical Equilibria in Soils. Wiley-Interscience, New York. Meeussen, J.C.L., Keizer, M.G. and De Haan, F.A.M., 1992a. The chemical stability and decomposition rate of iron cyanide complexes in soil solutions. Environ. Sci. Technol., 26( 3): 511-516. Meeussen, J.C.L., Keizer, MC., Van Riemsdijk, W.H. and De Haan, F.A.M., 1992b. Dissolution behavior of iron cyanide (Prussian blue) in contaminated soils. Environ. Sci. Technol., 26(9): 1832-1836. Meeussen, J.C.L., Keizer, MC., Van Riemsdijk, W.H. and De Haan, F.A.M., 1994. Solubihty of cyanide in contaminated soils. J. Environ. Qua]., 23(4): 785-792. Scheidegger, A., Biirgisser, Ch.S., Borkovec, M., Sticher, H., Meeussen, J.C.L. and Van Riemsdijk, W.H., 1994. Convective transport of acids and bases in porous media. Water Resour. Res., in press. Schwab, A.P. and Lindsay, W.L., 1983. Effect of redox on the solubility and availability of iron. Soil Sci. Sot. Am., 47: 201-205.
Geoderma67 ( 1995) 73-85
Transport of complexed cyanide in soil J.C.L. Meeussen *, W.H. van Riemsdijk, S.E.A.T.M. van der Zee Department of Soil Science and Plant Nutrition, Wageningen Agricultural University, P.O. Box 800.5. 6700 EC Wageningen, The Netherlands
Received 15 March 1994; accepted after revision 2 September 1994
Abstract Contamination and Fe(CN):-1,
of soil with cyanide, generally in the form of iron cyanide complexes [Fe( CN)iis commonly found at several types of industrial sites. The risks for human health
or the environment posed by such sites are largely determined by the chemical behaviour and transport of complexed cyanide in soil. In acidic soil this behaviour is probably dominated by equilibrium with Prussian blue, Fe,(Fe(CN),),(s), which is sparingly soluble under acidic conditions and limits dissolved cyanide concentrations and mobility. However, atpH levels higher than about 7 the solubility of this precipitate is extremely high, which allows cyanide to be very mobile under such conditions. Nevertheless, according to field observations, Prussian blue appears to persist for decades in alkaline soils. In this paper multicomponent transport calculations, including equilibrium with Prussian blue, iron hydroxide and relevant redox reactions, are compared with the situation as observed in two contaminated soil profiles. In case of an acidic soil, there appears to be good agreement between the calculations and the field situation. Here the mobility of cyanide is mainly regulated by the pH and redox potential and is relatively low. In case of an alkaline soil, precipitation of Prussian blue will not take place and cyanide will be very mobile once leached from the initial waste material. In this case the dissolution rate of Prussian blue is the process which determines cyanide concentrations in the groundwater. The calculations show that this dissolution rate is determined by the acid neutralizing capacity of the solution infiltrating in the waste material. The buffer capacity depends on the soil composition and can easily be so low that complete dissolution of Prussian blue may take decades. This very likely explains the persistence of this material in alkaline soils.
1. Introduction The use of cyanide in several kinds of industrial processes has regularly caused local contamination of the soil with this substance. It is commonly found as a soil contaminant * Corresponding author. Present address: The Macaulay Land Use Research Institute, Craigiebuckler, Aberdeen AB9 2QJ, Scotland, UK. 0016-7061/95/$9.50
0 1995 Elsevier Science B.V. All rights reserved
SsDIOOI6-7061(94)00061-1
14
J. C.L. Meeussen et al. / Geoderma 67 ( 1995) 73-85
at the sites of electroplating industries, metal ore processing industries and of coal gasification plants. Especially at the sites of (former) coal gasification plants, of which there are over 200 in the Netherlands, cyanide is often present in large amounts. These coal gasification plants have produced large amounts of cyanide containing waste material, which originated from the gas purification process. This waste material contained high concentrations of cyanide in the form of Prussian blue. Because this material was more or less worthless it was often stored or disposed of at the surface of the sites close to the gasification plants. In numerous occasions the presence of the waste material has caused contamination of the soil with cyanide by leaching of dissolved iron cyanide, or by contamination of the soil with the solid waste material itself. Because cyanide is a potentially highly toxic substance its presence in the soil may pose serious risks for human health and for the environment when it enters the ecological cycle by leaching to the groundwater. The risks for such hazards strongly depend on the bioavailability and mobility of cyanide in the soil. Both properties are strongly influenced by the chemical forms in which cyanide is present, and by the way these forms are bound to the soil solid phase. When cyanide was disposed of in the environment, in most cases it was discharged in the form of iron cyanide complexes [Fe(CN)iand Fe(CN)z- 1. These complexes are not very toxic (Fuller, 1984) but when exposed to daylight they readily decompose to highly toxic and volatile free cyanide [ HCN( aq) ] : Fe(CN)il-+6H,O+3H+%Fe(OH),(s)+6HCN(aq)
(1)
Free cyanide is not only toxic and volatile, but due to its uncharged nature also very mobile in soils ( Alesii and Fuller, 1976). Although free cyanide is susceptible to biodegradation in low concentrations, it is clearly the more hazardous species when compared to the complexed forms. According to thermodynamic equilibrium calculations, iron cyanide complexes tend to decompose to free cyanide in soils. Fortunately, this decomposition proceeds extremely slowly in darkness and thus in soils. In fact the decomposition rate of iron cyanide in soils is so low that it makes iron cyanide kinetically stable, and allows it to remain the predominant cyanide species in soils (Meeussen et al., 1992a). In case soil contamination has occurred with cyanide in the form of iron cyanide complexes, the behaviour of these complexes determines the overall behaviour of cyanide in the soil. This behaviour is governed by the interaction between dissolved iron cyanide complexes and the soil solid phase. In case of complexed cyanide the principal interaction with the soil solid phase is by precipitation and dissolution of Prussian blue, a mineral consisting of iron cyanide complexes and iron [ Fe,( Fe( CN),) S( s) ] (Meeussen et al., 1992b). The extra iron necessary for the formation of this mineral is generally abundantly present in soils, e.g. in the form of iron (hydr)oxides. Dissolution of Prussian blue can be represented by the following reaction: Fe4(Fe(CN),)-,(s)
%iFe(CN)i-
+4Fe,+
+3e-
(2)
Under aerobic conditions, the evolving Fe’ ?? will precipitate as an iron (hydr)oxide, e.g. Fe(OH),( s). This results in the following overall dissolution reaction for Prussian blue: Fe,(Fe(CN),),(s)
+ 12H,Or;i4Fe(OH),(s)
+3Fe(CN)i-
+3e-+
12H’
(3)
J. C. L. Meeussen et al. / Geoderma 67 (I 995) 73-85
75
According to this reaction, dissolution of Prussian blue consumes hydroxyl ions and produces electrons and therefore has an acidifying and reducing effect on the local environment. When Prussian blue dissolves under anaerobic conditions and little or no electron consuming agents are present, the evolving iron will partly remain in the reduced (Fe*+) state and more reduced states of iron hydroxide will form, e.g. Fe, (OH)*(s) (Schwab and Lindsay, 1983). This leads to the following overall dissolution reaction:
(4) The way that protons and electrons are involved in these dissolution reactions indicates that the solubility of Prussian blue depends on soil pH as well as on redox potential. In general a low pH and a low redox potential result in a low solubility of Prussian blue, and thus in low cyanide concentrations in the soil water phase and in a limited mobility of complexed cyanide. Additionally, precipitation or dissolution of Prussian blue may influence the pH and redox potential of the soil. The significance of this effect depends on the amount of Prussian blue present and on the transport behaviour of acid and base in the soil. The transport of acid and base in soils is governed by the buffering characteristics of the soil and the soil solution (Scheidegger et al., 1994). From the overall reaction equations it appears that four hydroxyl ions are needed to dissolve one mole of Fe(CN), from Prussian blue in case of reaction 3, and about 3.5 in case of reaction 4. This makes the solubility of Prussian blue extremely pH-dependent. According to simple equilibrium calculations, this precipitate would be completely soluble and thus mobile at the pH levels found in calcareous soils. Nevertheless, solid Prussian blue often appears to persist in such soils for at least decades after its disposal in waste material. In contrast with the situation in the acidic soils, where equilibrium calculations give reasonable predictions of occurring concentrations, the situations found in alkaline soils therefore cannot be explained by simply assuming a chemical equilibrium situation. The main aim of this work is to evaluate whether existing knowledge of the behaviour of complexed cyanide in soil and in the dissolution and precipitation of Prussian blue is adequate for modeling the transport of iron cyanide in contaminated soils, and to elucidate the cause for the persistence of Prussian blue in alkaline soil. This evaluation is made by comparing transport and accumulation of iron cyanide as observed at two contaminated sites to calculations with a multicomponent transport model in which the relevant chemical equilibrium reactions were incorporated. A further objective was to use this transport model for the identification of soil parameters that have a large impact on mobility of cyanide in soils. 2. Methods and materials 2. I. Field data
Field data on total cyanide content of the soil, pH, redox potential and dissolved cyanide concentrations were obtained by sampling two sites of former coal gasification plants (A
16
J.C. L. Meeussen et al. / Geoderma 67 (I 995) 73-85
and B). The collection of these field data has been described previously (Meeussen et al., 1994). The two sites were selected primarily because of their difference in soil type. Site A consists of an acidic sandy soil while at site B a calcareous clay soil is present. Both soil types are commonly prevalent in the Netherlands. The pH of the soil at the acidic site (A) was about 4.1, while the pH of the calcareous soil at site B was about 7.5. Because of the huge effect of pH on the dissolution and precipitation of Prussian blue, there is a substantial difference in behaviour of cyanide between these soils. At the surface of site A, the acidic site, no waste material in its original form remained at the time of the investigations. The top soil had apparently been replaced by clean soil material. At many contaminated sites the conditions are presently similar, especially when such sites are used for non-industrial purposes, e.g. for a residential area, which was the case here. The cyanide contamination of the soil present below the top soil layer was most likely caused by leaching of cyanide from waste material that previously covered the surface. The present soil consists of an aerobic layer that contained little cyanide and an anaerobic, water saturated, layer where cyanide had apparently accumulated. The accumulation was easily detectable due to the blue colour of this layer. On site B, the more alkaline site, the conditions were quite different. Here the original gasworks building was still present, although now used for other purposes. The surface of this site most likely has not changed much since gas production was discontinued here about thirty years ago. At least some of the original waste material was still present here, which was illustrated by clearly visible blue spots at the soil surface and by the high cyanide concentrations found in the top soil layer. In contrast with the acidic site, the blue colour was not observed in the soil layers below the surface. In the deeper layers there was virtually no cyanide present in the soil solid phase. However, high concentrations were found here in the soil water. 2.2. Calculations Calculations of transport and accumulation of iron cyanide were done using ECOSAT, a multicomponent transport computer model (Keizer et al., 1993). This model consists of a chemical module which calculates the speciation of chemical components and their distribution over different species and phases, and a transport module which calculates convective and dispersive transport of dissolved species. Both parts of the computer program are separated as much as possible. This allows the use of more sophisticated chemical or transport modules, and makes replacement of either of the modules by alternative ones relatively easy. This alternate calculation of chemical equilibrium and mass transport can cause extra numerical dispersion. However, in the ECOSAT program this dispersion is diminished by iteration between both modules until convergence is reached. The chemical reactions included in the model calculations are listed in Table 1. The most important assumption in modelling the transport of cyanide in this way is that interaction of dissolved iron cyanide with the soil solid phase only takes place by precipitation of Prussian blue (Table 1, reaction 3). Other possible interactions with the soil solid phase, for example adsorption of iron cyanide, were not considered. Another assumption was that in the acidic soil the buffer capacity for protons and electrons was completely supplied by dissolution and precipitation of iron (hydr) oxide minerals and of Prussian blue.
J.C.L. Meeussen et al. /Geodemm
67(1995)
73-85
71
Table 1 Equilibrium reactions and constants used in the transport calculations Reaction
log K”
Reference
3.54 4.63 - 84.5
Lindsay ( 1979) Lindsay ( 1979) Meeussen et al. ( 1992b)
13.0 6.3 - 2.2 -5.7 - 3.0 - 19.0 - 13.1 -21.6
Lindsay Lindsay Lindsay Lindsay Lindsay Lindsay Lindsay Lindsay
Solid phases
(1) Fe(OH),(s) +3H+ %Fe3++3H20 (2) Fe,(OH),(s)+8H+ S3Fe3++e-+8Hz (3) Fe,(Fe(CN),)3(s) %4Fe’* +3Fe(CN)zIron hydroxo complexes (4) Fe’+ +e-%Fe’+
(5) Fe3++H,0+e-%Fe(OH)+ +H+ (6) Fe3++HZOSFe(OH)*++Hf (7) Fe3++2H,0%Fe(OH):+2H’ (8) Fe3++2H,0+e-%Fe(OH)i+2Ht (9) Fe’++3H,O+e-%Fe(OH), +3H+ (10) Fe’+ +3H,O%Fe(OH)t+3H+ (11) Fe3++4H,0%Fe(OH),+4Ht Fe(CN), complexes (12) Fe(CN)z-+e-%Fe(CN)z-
(13) Fe(CN)i- +e-+H+SHFe(CN)i(14) Fe(CN)i-+e-+2H+ QH,Fe(CN)z-
+3e-
6.0 10.4 12.8
( 1979) (1979) ( 1979) ( 1979) ( 1979) ( 1979) ( 1979) ( 1979)
Beck (1987) Beck (1987) Beck (1987)
Carbonate equilibria
(15) (16) (17) (18)
CaC03(s) %Ca*+ +CO:CO:-+H+ %HCO, HCO; +H+ SH,CO; CO,(g) +HzO%H2CO;
-8.4 10.3 6.4 - 1.5
Lindsay Lindsay Lindsay Lindsay
( 1979) (1979) ( 1979) ( 1979)
In reality, of course, there are numerous other processes that may supply this kind of buffer capacity. The possible effects of this simplification are mentioned in the results and discussion section. In order to perform the calculations, the initial column composition with respect to total amounts of substances and initial pH of the solution was given as input. During the calculations the pH and redox potential were calculated, so these parameters were not fixed or manually adjusted. In case of the calculations with an acidic soil, the initial situation in the modelled column consisted of three separate zones. Although no longer present in reality, on top we located a layer of 0.1 m of waste material consisting of 0.01 M l- ’solid Prussian blue and 0.1 M 1- ’Fe( OH),( s), serving as source for cyanide leaching. Between this layer and the groundwater table we located an oxidized zone consisting of 0.1 M 1-i Fe(OH),( s) only, and below this layer a more reduced zone, where a mixture of 0.05 M l- ’ Fe( OH), (s) and 0.05 M I- ’Fe,( OH) s( s) was present. The amount of protons initially present in the soil was chosen to establish a pH of 4. The infiltrating solution consisted of water with a pH of 4 and did not contain cyanide. In the alkaline soil case, the initial situation in the modelled column was similar. Again three different zones were distinguished. A layer of 0.1 m of waste material, containing 0.01 mole Prussian blue l- 1 and 0.1 M I-’ Fe(OH),(s), was placed on top. Below this layer was located a homogeneous layer containing 0.1 M 1-l Fe( OH),( s). The deepest
78
J.C.L. Meeussen et al. / Geodena
67 (1995) 73%RS
layer consisted of a reduced zone where, like in the acidic soil, 0.05 M 1-l Fe(OH),( s) and 0.05 M 1-l Fe,(OH),( s) was present. In this case, the pH of the infiltrating solution was adjusted to 7.0, and the initial pH of the soil was adjusted to 7.2. In order to simulate the alkaline nature of the soil in this column, 0.1 M l- ’CaCO, in equilibrium with atmospheric CO, was added to the two layers below the contaminated layer. Without such a proton buffer the entire profile was likely to acidify within only a few pore volumes. As discussed below, calcite was not added directly to the contaminated layer because it is not stable in contact with Prussian blue. The transport calculations considered a soil column having a volumetric water content of 0.3 (water saturated), a water flux density of 0.35 m yr- ’ and for a period of 30 years of leaching.
3. Results and discussion 3. I. Acidic soil The cyanide content, pH and redox potential (or pe = 16.95&) as measured in the acidic contaminated soil are shown as a function of depth in Fig. 1, as are also the calculated values for these parameters after simulation of a thirty-year leaching period. The modelled situation shows a heavily contaminated top layer. If we first consider the pH profile, it Soil A measured ---- calculated
depth (cm) 0
-50
,_______..._...----.-.
I
j
$ -------
waste material
-100
-250
\
j
*
anaerobic zone
table 3
CN(t) soil mglkg Fig. 1. Profiles of total cyanide concentrations, calculated with the transport model, for soil A.
4
5
6
7
6
10
8
PH pH and redox potential as measured
12
pe in a field situation and as
J. C. L. Meeussen et al. / Geoderma 67 (1995) 73-85
79
appears that the modelled pH values are very close to the values observed in the field. The pH was initialized at 4 at the start of the calculations, but this is no guarantee for the pH to remain around this value during the calculations. The calculations show a slight pH increase in the aerobic zone ( f0.2 units) compared to the initial pH of 4. This increase occurs during the first few pore volumes of leaching and is the result of the precipitation of a small amount of Prussian blue. Hereafter, in this zone the pH remains constant and no further precipitation of Prussian blue occurs. At the top of the anaerobic layer a more profound increase of the pH can be observed ( f 0.5 units). A comparable increase is observed in the field situation. Another relevant parameter is the redox potential of the soil solution. As also observed in the case of pH, it appears that the modelled results agree quite well with the field measured values, in the aerobic as well as in the anaerobic zone. In general, correct determination of redox potentials in soil solutions is difficult, and observed values are usually not very reliable. However, determination of the redox potential is relatively reliable in solutions containing Fe(CN), ions, due to the fast redox equilibrium between Fe(CN)i- and Fe(CN)z-. The calculated pH and redox values are sufficiently close to the real values to allow prediction of the location of dissolution of Prussian blue in the soil, although the pH and redox buffering mechanisms in the real soil are without doubt much more complex than the simple precipitation reactions of the different minerals as implemented in the model. In the field situation, iron cyanide appears not to accumulate significantly in the aerobic zone. Hence, only very small amounts of iron cyanide were found here in the soil solid phase. The same outcome was observed in the modelled results. During the calculations the initial conditions in the column favoured precipitation of Prussian blue in the aerobic zone and initially some Prussian blue precipitated there. However, this resulted in an increased pH which prevented further precipitation. In this way, only a small part of the leached iron cyanide precipitated and became immobile while the rest of the iron cyanide remained available for transport and thus was mobile in this layer. In the calculations, the exact amount of iron cyanide that precipitates in this layer is determined by the initial pH of the soil solution. In the field this amount is strongly influenced by the soil’s proton buffer capacity. If this buffer capacity is large enough, precipitation of Prussian blue does not affect pH, and complete precipitation may occur. In the anaerobic zone, especially in the upper part of it, accumulation of iron cyanide is observed both in the field situation and in the calculations. In this layer the solubility of Prussian blue is lower than in the aerobic zone because of the lower redox potential. The Prussian blue that precipitates here tends to increase the redox potential and the pH. In the model, this effect is neutralized by the transformation of Fe,(OH)8 to Fe( 0H)3, which buffers the redox potential. Therefore, the solubility of Prussian blue remains low and accumulation is possible in this layer. When the conversion of Fe,(OH), to Fe(OH), is complete, no further precipitation of Prussian blue occurs. It is possible that in the field situation the same processes buffer the pH and redox conditions, but it is very likely that here organic matter will also act as a reducing agent, making continued precipitation of iron cyanide possible. The layer where accumulation of cyanide occurs in the field is more diffuse than calculated with the model. This is probably mainly caused by fluctuations in the groundwater table in the field, which causes the redox profile here to be more diffuse than in the model calculations.
80
J.C.L. Meeussen et al. /Geodertna 67 (1995) 73-B
Soil B depth (cm) ---
measured calculated
I:
._-_-__.-__._.__-._.-
-200
groundLater table I I 1
-250
1
500
CN(t) soil mglkg
1000
3
4
5
PH
6
7
6
8
10
12
pe
Fig. 2. Profiles of total cyanide concentrations, pH and redox potential as measured in the alkaline field situation and as calculated with the transport model, for soil B.
3.2. Alkaline soil The observed situation in the alkaline field soil differs in several aspects from that in the acidic soil. First, the highest concentration of soil cyanide was here present in the top soil layer, which most likely contained some original waste material. The blue colour of this layer strongly suggested the presence of Prussian blue. In the deeper layers, where much less cyanide was present, this colour was absent. Most of the cyanide in these layers was found in the dissolved phase (Meeussen et al., 1994). The concentration of iron cyanide measured in the soil water of the saturated zone at this site was more than ten times higher thaninthesoilwateroftheacidicsite[1.1X10~5MFe(CN),comparedto9.3X10~7M Fe( CN),] . When we look at the pH profile of the soil (Fig. 2) it appears that the pH at this site is significantly higher than at the acidic site. There is also a more pronounced difference in pH between the layers, with pH of the top soil about 5.6, while in the deeper layers it is about 7.2. 3.3. Modelling results Let us first consider the results of the calculations with the alkaline soil, which had an initial pH of the column of 7.2, and the addition of calcite to the subsurface soil layers to buffer the pH there. Immediately after the calculations were started, the pH in the top layer decreased to about 5.2, due to the dissolution of Prussian blue. A very similar pH increase with depth was observed in the field, although pH in the upper soil layer was slightly higher here.
J. C. L Meeussen et al. / Geoderma 67 (I 995) 73-%5
81
The pH of the leachate was buffered to about 7.3 by dissolving calcite in the deeper layers. The dissolved iron cyanide that leached from the contaminated layer did not reprecipitate at the alkaline pH levels in the deeper layers and remained mobile. Dissolution of Prussian blue in alkaline soils is clearly a one direction process. This is largely in agreement with the field observations where precipitated iron cyanide was also present only in the contaminated top layer. In these calculations, the concentrations of dissolved iron cyanide which can occur in the ground water depend mainly on the dissolution rate of Prussian blue in the waste material (upper layer). The acidification of this layer limits the solubility of iron cyanide in the calculations to about 5 X 10e6 M Fe(CN)6 in case of leaching with pure water, which is about the same concentration as observed and calculated in the case of the acidic soil. However, the iron cyanide concentrations found in the alkaline soil water are well above this concentration. A possible explanation for this phenomenon follows. One of the main reasons for the higher pH at the alkaline site is the occurrence of calcite. Dissolution of this mineral buffers the pH at an alkaline level. Calcite can also react with Prussian blue according to the following reaction: Fe,(Fe(CN),),(s) *4Fe(OH),(s)
+6CaCO,(s) +3Fe(CN)z-
+ 12H,O +6H2C03+6Ca2+
(5)
According to chemical equilibrium calculations, this acid base reaction proceeds completely, which implies that Prussian blue and calcite are not chemically stable when mixed. One of the two minerals will dissolve entirely, depending on which of them is more abundant. In that case, the resulting concentration of dissolved iron cyanide is equal to half the concentration of dissolved calcium. An alkaline pH will develop when there is an excess of calcite, while an acidic pH will evolve when there is more Prussian blue present. According to the acidic pH of the top soil layer, calcite is most likely absent here, while according to the measurements and the soil colour Prussian blue is absent in the deeper layers. When solid Prussian blue is spatially separated from solid calcite, the reaction between both minerals, and thus dissolution of both, can still proceed via reaction with mobile dissolved ions. When calcite reacts with dissolved carbon dioxide it dissolves according to the following reaction: CaCO, ( s) + H.&O, * 2HCO; + Ca*+
(6)
The evolving dissolved ions are able to migrate to solid Prussian blue, where the following reaction takes place: Fe,(Fe(CN)&(s)
+ 12HCO; + 12H,O%4Fe(OH),(s)
+3Fe(CN)z-+6H,CO, (7)
The combination of reaction 6 (6 times) with reaction 7 is equivalent to reaction 5. The maximum rate at which this reaction proceeds is, apart from pure chemical dissolution kinetics, limited by the rate at which especially the HCO; ion is transported from one mineral to the other. In case of short distances between these solid materials, this transport rate will be governed by diffusion. An expression for the flux density in case of diffusion is given by Fick’s law for molecular diffusion, with modifications for porosity and pore tortuosity:
82
J. CL. Meeussen et al. / Geoderm
67 (1995)
J= 8.D.(Sc/Sx)
73-85
(8)
where J is the flux density (in mol m-2 s- ’), 8 is the porosity, D is the diffusion coefficient in soil, which is derived from the molecular diffusion coefficient in free solution after correction for the tortuosity, which has a value of about 1.25 -9 m2 s- ‘, c is the concentration (in mol m- “), and x the distance (in m) (Bolt, 1982). The factor that determines the transport rate in this case is the gradient of ion concentration. Because the concentration differences are relatively constant due to buffering by the dissolving minerals, the diffusion rate will be predominantly determined by the distance between the solid phases. In case of short distances between both solids, e.g. when both minerals are present in very small particles and are intensively mixed, these diffusion fluxes are high, and allow fast dissolution of both minerals. However, under such conditions dissolution rates might be limited by pure chemical dissolution kinetics. Apart from diffusion, the transport of dissolved ions can also result from convection. In humid climates this is the dominant transport process. In case of convective transport the flux density equation equals: J= 8.11.~
(9a)
where J is the flux density (in mol rnp2 s- ’), 19is the porosity, c is the concentration (in mol m- 3), and u is the the pore water velocity (in m s - 1) . However, the effective convective flux of dissolved ions to (or from) a solid particle is determined by the difference in concentration between the solution before and after contact with the particle. In case of instantaneous equilibrium this is the difference between the concentration in the bulk solution and the concentration in equilibrium with the solid phase. If we consider this concentration difference to be 6c, this leads to: J= e.u.8~
(9b)
By combining Eqs. (8) and (9a) it is possible to estimate the distance x, at which convective transport becomes predominant over diffusive transport. From the combination of both equations follows the expression for the condition where the diffusive flux equals the effective convective flux: e.D.(6c/Sx)
=e.Lj,sc
( 10)
If we consider the distance to be x this can be reduced to:
(11) The net infiltration rate in the Netherlands on average equals 0.35 m yr- ‘, which results for water saturated conditions in a pore water velocity u of 3.6 X lo-’ m s- ’) . The unknown distance x is found by using Eq. ( 11) which results in 0.035 m. This implies that convective transport of ions dominates when solid calcite is separated more than about 3.5 centimetres from solid Prussian blue. This calculation is based on an averaged water velocity. However, convective water fluxes during wetter periods in the field can be orders of magnitudes higher, which makes convective transport already dominant at much shorter distances (Bolt, 1982).
J.C.L. Meeussen etal. / Geodemza 67 (I 995) 7345
83
0
-2 -
-3 -
-4 -
-5 -
-6
I -7
I -6
I -5
I -4
log alkalinity
I -3
I -2
I -1
I 0
(mol/l)
Fig. 3. Calculated concentration of dissolved cyanide in leachate from Prussian blue containing waste material as function of the alkalinity of the infiltrating solution.
Because original waste material was not well mixed with the soil, convective transport of alkalinity will in most cases be the dominating process regulating the dissolution of Prussian blue. In that case the concentration of iron cyanide that results from dissolution depends on the capacity of the infiltrating solution to neutralize acid. The acid neutralizing capacity is determined by the alkalinity of the solution, which equals [ OH - ] + [ HCO; ] + 2 [CO: - 1. If we assume that an infiltrating solution equilibrates with Prussian blue while passing a contaminated layer, the resulting iron cyanide concentrations can easily be calculated with the ECOSAT model (Fig. 3). At high alkalinities, the concentration of Fe( CN)z- in the leachate is equal to four times the alkalinity because four moles of alkalinity are needed to dissolve one mole of Fe(CN)z-. At low alkalinities, the minimum concentration of dissolved iron cyanide is limited by the solubility of Prussian blue in pure water (which has an alkalinity of zero). The concentrations in Fig. 3 represent maximum concentrations, because in the field heterogeneity can cause incomplete contact between leaching water and Prussian blue. Nevertheless, with these assumptions a maximum value for iron cyanide concentrations in the leachate of cyanide-containing waste material can be derived. In calcareous soils, the alkalinity of the soil water, is strongly influenced by equilibrium between calcite and atmospheric carbon dioxide, and ranges from 1O-4 to 10e3 M (Lindsay, 1979). If this soil water leaches through the Prussian blue containing layer, iron cyanide concentrations of about 4 X lop5 M to 2 X 10e4 M will result, which are higher than observed in the field. If there is no equilibrium with calcite in the upper layer, the alkalinity of soil solution in the top soil layer is probably significantly lower. If we assume an alkalinity of low5 M, for which a CO2 pressure of 3 mbar is needed at the pH of the top soil, this results in concentrations of dissolved cyanide which are in the range of the concentrations as observed in the
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field. Differences between observations and calculations can very well be caused by mixing of contaminated groundwater with clean groundwater or because either not all of the infiltrating water is in equilibrium with calcite or it equilibrates with Prussian blue while leaching. Each of these factors may lower the concentration of iron cyanide in leachates. Nevertheless, these assumptions can be used to estimate the time necessary to dissolve all Prussian blue in this soil. When we consider the present situation at site B, it is possible to estimate the time it will take before dissolution is complete if dissolution rates remain constant in time. At site B, a layer of about 10 cm thickness with a total cyanide content of 930 mg CN kg- ’is present. This is equal to 6 X 10M3 M Fe( CN), kg- ’which makes about 1 X 10-2MFe(CN)61-’ at anestimatedbulkdensityof 1500 kg m-‘. Underthe hydrologic conditions in the Netherlands, with a net water flux of 0.35 m yr~ ’ and a concentration in theleachateof1.1X10~5,itwouldtake(1X10~2)/(1.1X10~5X3)=300year.Thisis clearly long enough to justify the survival of Prussian blue in an alkaline soil for over thirty years. At the calculated maximum leaching concentration of 2 X 10e4 M l- ’ complete dissolution of the cyanide would take about 15 years.
4. Conclusions
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-
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Observed accumulation of iron cyanide in the field qualitatively agrees with transport calculations based on the assumption that interaction of dissolved cyanide with the soil solid phase is dominated by precipitation of Prussian blue. In soils with pH levels below about 5, concentrations of dissolved iron cyanide are governed by equilibrium with Prussian blue. Prussian blue is not stable and will eventually dissolve in alkaline soils. Here, dissolution rates of Prussian blue and the resulting concentrations of dissolved iron cyanide will be largely governed by the alkalinity of the infiltrating solution. To leach iron cyanide from a contaminated soil for remediation purposes, a solution with a high alkalinity is needed. Under field conditions leaching rates in alkaline soils can easily be so low that complete leaching of iron cyanide from contaminated soils can take decades.
References Al&i, B.A. and Fuller, W.H., 1976. The mobility of three cyanide forms in soils. EPA-600/9-76-015, Residue Management and Land Disposal, Proc. Hazard. Waste Res. Symp., pp. 213-223. Beck, M.T., 1987. Critical evaluation of stability constants of metal complexes in solution. Critical survey of cyano complexes. Pure Appl. Chem., 59: 1703-1720. Bolt, G.H., 1982. Soil Chemistry, B. Physic0 Chemical Models. Elsevier, Amsterdam. Fuller, W.H., 1984. Cyanide in the environment with particular attention to Soil. In: D. Van Zyl (Editor), Cyanide and the Environment, I. Geochemical Engineering Program. Colorado State Univ., Fort Collins, pp. 1946. Keizer,M.G., De Wit, J.C.M., Meeussen, J.C.L.,Bosma, W.J.P.,Nederlof, M.M., Venema,P., Meeussen, V.C.S., Van Riemsdijk, W.H and Van der Zee, S.E.A.T.M., 1993. ECOSAT, A computer program for the calculation
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of chemical speciation and transport in soil-water systems. Dep. of Soil Science and Plant Nutrition, Wageningen Agricultural University, Wageningen. Lindsay, W.L., 1979. Chemical Equilibria in Soils. Wiley-Interscience, New York. Meeussen, J.C.L., Keizer, M.G. and De Haan, F.A.M., 1992a. The chemical stability and decomposition rate of iron cyanide complexes in soil solutions. Environ. Sci. Technol., 26( 3): 511-516. Meeussen, J.C.L., Keizer, MC., Van Riemsdijk, W.H. and De Haan, F.A.M., 1992b. Dissolution behavior of iron cyanide (Prussian blue) in contaminated soils. Environ. Sci. Technol., 26(9): 1832-1836. Meeussen, J.C.L., Keizer, MC., Van Riemsdijk, W.H. and De Haan, F.A.M., 1994. Solubihty of cyanide in contaminated soils. J. Environ. Qua]., 23(4): 785-792. Scheidegger, A., Biirgisser, Ch.S., Borkovec, M., Sticher, H., Meeussen, J.C.L. and Van Riemsdijk, W.H., 1994. Convective transport of acids and bases in porous media. Water Resour. Res., in press. Schwab, A.P. and Lindsay, W.L., 1983. Effect of redox on the solubility and availability of iron. Soil Sci. Sot. Am., 47: 201-205.