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Geochemical changes during biodegradation of petroleum hydrocarbons: field investigations and biogeochemical modelling
Organic Geochemistry 30 (1999) 423±435
Geochemical changes during biodegradation of petroleum hydrocarbons: ®eld investigations and biogeochemical modelling H. Prommer a, b,*, G.B. Davis b, D.A. Barry c a
Department of Environmental Engineering, University of Western Australia, Nedlands, 6907, Australia b Centre for Groundwater Studies, CSIRO Land and Water, Wembley, 6014, Australia c School of Civil and Environmental Engineering and Contaminated Land Assessment and Remediation Research Centre, University of Edinburgh, Edinburgh, EH9 3JN, UK
Abstract Sediment analysis using wet extraction techniques and model simulations were carried out to investigate the role of ferric iron during biodegradation of dissolved petroleum hydrocarbon compounds at a ®eld site in Perth, Western Australia. Sediment cores were analysed for iron and sulphur species. The total iron concentrations were found to be low while the fraction of Fe(II) was surprisingly high. Pyrite was detected in the contaminated zone, but no iron monosulphides were found. A reactive multi-component transport model, coupling advective-dispersive transport of organic compounds and inorganic aqueous components with a geochemical equilibrium model and a biodegradation module, was applied to simulate qualitatively the ¯uxes and reactions involved in the biodegradation of BTEX compounds. Both ®eld investigations and modelling suggest that ferric iron minerals play no important role as electron acceptors while sulphate provides the major part of the oxidation capacity. # 1999 Elsevier Science Ltd. All rights reserved. Keywords: Groundwater; Iron reduction; Sulphate reduction; Multi-component transport; PHREEQC; BTEX
1. Introduction Petroleum products containing benzene, toluene, ethylbenzene and xylene (BTEX) have contaminated many shallow aquifers (Barker et al., 1987; Schmitt et al., 1996). Degradation of these compounds, catalysed by bacteria, has been shown to occur at dierent rates in many geochemical environments. BTEX degradation is well known in aerobic environments but is also reported to occur under nitrate-reducing (Zeyer et al.,
* Corresponding author. Fax: +61-8-9380-1015. E-mail address: [email protected] (H. Prommer)
1986; Kuhn et al., 1986; Barbaro et al., 1992), sulphate-reducing (Edwards et al., 1992; Beller et al., 1992; Wisotzky and Eckert, 1997), iron-reducing (Lovley et al., 1989; Lovley and Lonergan, 1990) and methane-producing (Grbic-Galic and Vogel, 1987) conditions. Generally, in conditions where neither the carbon-source nor nutrients limit microbial growth, the microbial communities consuming the thermodynamically most favourable electron acceptor outcompete other communities until the availability of the electron acceptor becomes limited. Thus, electron acceptors appear to be consumed sequentially, forming discrete redox-zones (Baedecker et al., 1993). However, in cases where both sulphate and Fe(III) serve as terminal elec-
0146-6380/99/$ - see front matter # 1999 Elsevier Science Ltd. All rights reserved. PII: S 0 1 4 6 - 6 3 8 0 ( 9 9 ) 0 0 0 2 7 - 3
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tron acceptors, the redox-sequence is not always clear. Lovley and Phillips (1987b) found inhibition of sulphate reduction by ongoing ferric iron reduction whereas Ludvigsen et al. (1995), for example, reported ongoing sulphate reduction in a land®ll rich in reducible iron sediments. Our study is aimed at identifying the role of ferric iron in the degradation process of BTEX at a particular contaminated shallow sand aquifer beneath Perth/Western Australia. More speci®cally we quantify the contribution of iron as a potential electron acceptor. If Fe(III) is not involved, then degradation at this ®eld site relies solely on sulphate as the electron acceptor implying that the observed degradation rates (Davis et al., 1999) are associated with sulphate reduction. 2. Field site The ®eld site is located in the Perth Metropolitan area. A leaking underground storage tank (UST) released approximately 500 l of gasoline, leading to a plume of more than 400 m in length for some compounds (see Figs. 1 and 2). Organic compounds and selected inorganic species have been monitored at the ®eld site since 1991 at up to 18 multiport boreholes with vertical spacings of 0.20±0.5 m (see Davis et al., 1999 for details). Biodegradation was shown to occur and (local) degradation rates were determined in situ via tracer tests using fully deuterated benzene, toluene, p-xylene and naphthalene (Thierrin et al., 1995). A detailed description of the contamination and its temporal and spatial changes is given by Davis et al. (1999). 2.1. Hydrogeology The site is part of the quaternary Bassendean Sand aquifer. At the ®eld site, the aquifer is 7 to 12 m thick and is underlain by a clay aquitard (Barber et al., 1991). The sands are generally medium-to-®ne grained bleached quartz sands. With the zone of water table ¯uctuation, organic matter has accumulated and the sands form a dark-brown, variably cemented 0.05±0.6m thick layer (Davis et al., 1999), locally called `coee rock'. The groundwater ¯ow direction changes seasonally by up to 208 while the water-table elevation varies by up to 1.8 m. Groundwater velocity at the ®eld site is estimated to be between 100 and 170 m/yr (Davis et al., 1999). 2.2. Local geochemistry The background geochemistry (see Table 1) of the contaminated site is generally anaerobic. Furthermore, nitrate concentrations are very low (0.1±1 mg/l,
Gerritse et al., 1990) though the residential area upstream of the study site is unsewered. Intensive denitri®cation is assumed to be the main reason for the low nitrate levels. In contrast to nitrate reduction, sulphate reduction is not thought to occur naturally in the Bassendean Sands (Gerritse et al., 1990). Thus, sulphate provides the largest potential oxidation capacity of the uncontaminated groundwater at the ®eld site. Typically, sulphate concentrations measured at MP9 (Fig. 1), a background multiport bore upstream of the underground storage tank, were between 50 and 100 mg/l. The average pH of background groundwater generally lies between 5 and 5.7. It decreases approximately 0.5 units between the top port and the bottom port of bore MP9, as can be seen in Fig. 2. Within the BTEX-contaminated zone, sulphate concentrations were signi®cantly lower than outside the plume. This can be seen on the contour plot for August 1994 (Fig. 1), in the appropriate depth pro®les along the main ¯owline (Fig. 2) and in Table 1, where mean values of aqueous concentrations (organic and inorganic compounds) outside and within the plume are listed for comparison. The lowest sulphate concentrations were generally found close to the contamination source (MP12, MP11). Hydrogen sulphide was found in low concentrations (<3 mg/l) in samples from within the contaminant plume. Bicarbonate concentrations were signi®cantly higher within the plume whereas no clear trend was exhibited for Fe(II). Despite the minor dierence in mean values (Table 1), signi®cantly increased Fe(II) concentrations (compared to background values) were occasionally (especially in the ®rst few years after the BTEX spill) measured in some ports of wells closer to the contamination source. In contrast, Fe(II) concentrations further downstream, e.g. at MP6, seemed to be lower within the contaminated zone (Fig. 2). The pH increased typically by 0.2±0.7 units within the plume while measured Eh values, as expected, decreased (Table 1). 3. Sediment analysis 3.1. Methodology 3.1.1. Sampling For this study, two cores were recovered from the ®eld site in November 1996. One was taken from the uncontaminated zone up-gradient of the spill halfway between the spill (UST on Fig. 1) and bore MP9. It was thought to provide background concentrations. The other core was recovered from the contaminated zone, 80 m downstream from the spill area next to bore MP1. As the BTEX plume is relatively thin, it was expected that the core from the contaminated
Fig. 1. Benzene, toluene and sulphate concentrations at the ®eld site in August 1994 (Davis et al., 1995),
H. Prommer et al. / Organic Geochemistry 30 (1999) 423±435 425
Fig. 2. Depth pro®les of selected organic compounds and inorganic species in August 1994; Depth=depth below water table.
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H. Prommer et al. / Organic Geochemistry 30 (1999) 423±435 Table 1 Mean values for inorganic compounds and physical parameters inside and outside of the contaminant plume (Linge, 1996) Parameter
Meanbackgr.a
Meaninsideb
O2 (mg/l) NOÿ 3 (mg/l) NH+ 4 (mg/l) Fe2+ (mg/l) SO2ÿ (mg/l) 4 H2S (mg/l) HCOÿ 3 (mg/l) pH Eh (mV)
0.56 0.18 0.25 1.6 82.5 0.1 14.3 5.1 103
0.16 0.05 0.3 1.4 32.4 1.1 74 5.7 ÿ 30
a
Total BTEX concentration R1 mg/l, 102 samples, except + for NOÿ 3 and NH4 (30 samples) and H2S (20 samples). b Total BTEX concentration >1 mg/l, 170 samples, except + for NOÿ 3 and NH4 (80 samples) and H2S (54 samples).
zone would show a pronounced depletion of reducible iron in the vertical pro®le. The cores were recovered in aluminium tubing (5-cm diameter) using a hollow-stem auger drilling rig. The aluminium tubing was cut into 25-cm sections, sealed immediately and placed within an anaerobic chamber within several hours. At the time of sampling, the groundwater table of both cores was located between the top section (25-cm section) and the second section. The bottom sections of both cores are taken from a depth of approximately 2.5 m below the appropriate groundwater table. The top 10 cm and bottom 10 cm of each section was discarded to minimise the in¯uence of any reoxidation taking place at the core ends. The remaining sediment material was then mixed thoroughly and subsamples of 2±3 g were placed in 20-ml glass serum bottles. 3.1.2. Analytical procedure The sediments were analysed and concentrations quanti®ed using the wet extraction techniques as tested and described by Heron et al. (1994). Extractions using 1 M CaCl, 0.5 M HCl and 5 M HCl were carried out, corresponding to ion-exchangeable, amorphous and crystalline forms of iron, respectively. After addition of the appropriate extractant (10 ml) in the anaerobic chamber, the glass bottles were sealed with a rubber stopper. Samples tested for ion-exchangeable iron were rotated for 8 h before being analysed for Fe(II) and Fe(III). The extraction time of the 0.5 M HCl extraction was 24 h. The 5 M HCl-extracted samples were heated for 1 h at 1208C. This procedure (Linge, 1996) modi®es slightly that of Heron et al. (1994) and was found to dissolve iron more completely. In particular, the method was able to completely extract even Fe(II) carbonates. For each sampling depth, triplicates were
427
analysed for Fe(II) and the total iron concentration (Fetot) by a colorimetric method using ferrozine; absorbance was determined with a VARIAN 100 DMS spectrophotometer. Quanti®cation of total iron concentrations was carried out by reducing iron completely to Fe(II) with hydroxylamine hydrochloride. Fe(III) concentrations were obtained from the dierence between Fetot and Fe(II). AVS (acid volatile sulphur) was analysed to determine FeS concentrations (Heron et al., 1994), assuming that AVS was all FeS. A sequential extraction by HI, and a 0.7 M Cr(II), 2 M HCl mixture was used for FeS2. Extractions were carried out using a digestion apparatus in which the H2S released was trapped in a zinc acetate solution. The methylene blue method (Cline, 1969) was used to quantify the extracted sulphides. Pyrite concentrations were then calculated from the dierence between HI-extracted sulphides and the Cr(II)-extracted sulphides. 3.2. Results The methodology described above resolved the depth pro®les of iron in the two cores taken. The errors within the triplicate samples typically were smaller than the dierences between two subsequent sampling depths. However, concentrations of ionexchangeable, amorphous and crystalline iron in both cores were lower than initially expected. A maximum total iron concentration of 900 mgFe/gsoil was found by 5 M HCl extraction for the yellow sands at the top of the background core (Fig. 3), recovered outside the BTEX plume. Below the yellow sand, two samples between 17 and 17.5 m AHD showed a bleached, white sand with minimal iron concentrations. Below these white sands, two samples showed dark brown `coee rock' with concentrations above 200 mgFe/gsoil of 5 M HCl-extractable iron. Concentrations less than 200 mgFe/gsoil were found in samples of light brown sand below the `coee rock'. In the core taken from the contaminated zone, the top four subsamples (16.5 and 17.5 m AHD) consisted of bleached white sand followed by one subsample of dark brown `coee rock' (approx. 16.25 m AHD), having the maximum total iron concentration of that core (200 mgFe/gsoil). Below that depth, total iron concentrations dropped back steadily as the sand became more and more bleached. From the total iron concentrations found, the largest fraction seemed to be Fe(II), even in the background core. This indicates that a large part of the reactive Fe(III) oxides were reduced before the aquifer was contaminated by BTEX, presumably by oxidation of natural organic matter. Reducible iron was mainly found as 5 M HCl-extractable iron in the background core, suggesting that Fe(III) minerals mainly occur in crystalline form. The concentrations were almost all
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Fig. 3. Concentrations of 5 M HCl and of 0.5 M HCl-extractable iron in units of m/g of soil; values for Fe(III) were calculated from Fetot ÿ Fe(II); AHD=Australian Height Datum.
around 40±60 mgFe/gsoil (see Fig. 3). Ion-exchangeable iron concentrations (not shown) were insigni®cant in all samples. Iron monosulphide concentrations (FeS, concentrations not shown) were not signi®cant in either core. Pyrite was present in 5 of 10 samples in the core taken
from the contaminated zone, but in none of the samples from the background core. However, the sequential HI/Cr(II) attack showed a high variability amongst the triplicates analysed, especially compared to the other extraction methods. In all background core samples and in some locations in the core from
H. Prommer et al. / Organic Geochemistry 30 (1999) 423±435
the contaminated area, HI-extracted sulphide concentrations were higher than the concentrations extracted by Cr(II) attack (Fig. 4).
4. Electron balance An electron balance calculation was carried out to verify whether the amount of Fe(III) found in the analysis would be sucient to account for the observed degradation of the BTEX compounds. We therefore assume that all Fe(III), including the fraction extracted by 5 M HCl, is available for the oxidation of the organic contaminants. This assumption is supported by observations of aquifer sediments contaminated by leachate from a Danish land®ll (Heron and Christensen, 1995). There, it was suggested that, over time, most of the crystalline Fe(III) oxide content can be utilised for organic matter degradation. In addition, Roden and Zachara (1996) noted that natural crystalline oxides may function as more eective electron acceptors than has been recognised previously. For the electron balance, we consider only electrondonating reactions of BTEX compounds, e.g. for toluene: ÿ C7 H8 21H2 O47COÿ2 3 50H 36e ,
1
while, for the electron-accepting reaction, we consider reduction of goethite: FeOOH eÿ 3H 4Fe2 2H2 O:
429
lents/m3) were available for the BTEX degradation, a bulk volume of roughly 44,300 m3 would be needed to achieve complete mineralisation of the BTEX compounds by iron reduction only. Considering the seasonal variability of the groundwater ¯ow direction, this is in fact approximately the bulk volume in contact with the BTEX plume. A rough estimate for an annual electron ¯ow, Eas, due to sulphate reduction, ÿ ÿ SOÿ2 4 9H 8e 4HS 4H2 O,
3
neglecting dispersive ¯uxes, might be derived from: Eas vne Ap DCSOÿ2 nt , 4
4
where v is the pore water velocity, ne is the (eective) porosity, Ap a cross-sectional area, nt is the number of electrons transferred and DCSOÿ2 is the average change 4 in sulphate concentration due to sulphate reduction within water passing through the cross-sectional area Ap. With groundwater ¯ow velocities of 100±170 m/yr, ne and nt of 0.3 and 8, respectively, and estimates for Ap and DCSOÿ2 of 60 m2 and 20 mg/l respectively, 4 between 4000 and 6700 equivalents are consumed per year. The latter estimate indicates that iron-reduction certainly cannot be the only electron-accepting process. However, it becomes also clear that, despite the low concentrations of Fe(III) oxides, there was potential for a signi®cant contribution of iron as an electron acceptor, if it could locally outcompete sulphate reduction.
2
Neglecting any biodegradation, vaporisation or other processes reducing the total contaminant mass while it passes through the unsaturated zone, we might estimate that 1600 mol of BTEX or approximately 57,600 equivalents (using toluene as a model compound) have reached the saturated zone. Assuming that 40 mgFe(III)/gsoil (=1.3 mol of Fe(III)/m3 or 1.3 equiva-
5. Geochemical equilibrium considerations As pointed out by Postma and Jakobsen (1996), microbially mediated degradation reactions can be seen as a two-step process, a ®rst step being the fermentation of, e.g. organic matter and the second being a subsequent consumption of the fermentation products
Fig. 4. Concentrations of sulphur in units of mg/g of soil (lines); Cr(II) extraction=D, HI=w; AHD=Australian Height Datum.
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H. Prommer et al. / Organic Geochemistry 30 (1999) 423±435
Table 2 Measured and equilibrated aqueous and mineral concentrations Aqueous component,a Mineral
Cbackgr. measured (mol/l)b
Cbackgr., Cin¯ow equilibrated (mol/l)b, (S1)
Cbackgr., Cin¯ow equilibrated (mol/l)b, (S2)
Cbackgr., Cin¯ow equilibrated (mol/l)b, (S3)
pH pc Eh (mV) C(IV) Alkalinity S(VI) S(ÿII) Fe(II) Fe(III) Ca Mg Na K Cl Fe(III) Fe(II) Goethitec Siderite Pyrite Magnetite
5.0 ± 129
5.21 ÿ 0.45 ± 3.69 10ÿ2 2.97 10ÿ3 7.84 10ÿ4 5.27 10ÿ12 1.33 10ÿ3 1.61 10ÿ12 8.41 10ÿ4 5.66 10ÿ4 5.24 10ÿ3 2.59 10ÿ4 6.45 10ÿ3
5.07 1.21 ± 7.95 10ÿ3 4.29 10ÿ4 7.84 10ÿ4 9.27 10ÿ24 6.65 10ÿ5 2.14 10ÿ12 8.41 10ÿ4 5.66 10ÿ4 5.24 10ÿ3 2.59 10ÿ4 6.45 10ÿ3
5.07 1.21 ± 7.95 10ÿ3 4.29 10ÿ4 7.84 10ÿ4 9.27 10ÿ24 6.65 10ÿ5 2.14 10ÿ12 8.41 10ÿ4 5.66 10ÿ4 5.24 10ÿ3 2.59 10ÿ4 6.45 10ÿ3
1.00 10ÿ3 4.00 10ÿ3 0.0 0.0
1.00 10ÿ3 0.0 0.0 0.0
1.00 10ÿ3 ± 0.0 4.00 10ÿ3
3.27 10ÿ4 7.83 10ÿ4 < 3.0 10ÿ6 6.65 10ÿ5 ± 8.41 10ÿ4 5.65 10ÿ4 5.23 10ÿ3 2.58 10ÿ4 6.44 10ÿ3 0.4±1.2 10ÿ3 2.0±4.0 10ÿ3 0.0
a
Values in parentheses indicate valence. If not indicated otherwise. c Representing crystalline iron.
b
(e.g. acetate, formate, H2) by the dierent terminal electron-accepting processes (TEAPs). Postma and Jakobsen (1996) also argue that, if fermentation is the rate-controlling step in the overall degradation reaction, the reactions can be described as a partial equilibrium process (Zehnder and Stumm, 1988) in which the electron-accepting process is an equilibrium reaction. They have used that approach for geochemical equilibrium considerations at the Fe(III)/sulfate-reduction interface and show that, in dierent geochemical environments (depending on the stability of the Fe(III) oxides and the pH), either sulphate reduction or iron reduction could be thermodynamically more favourable. As in other models (McNab and Narasimhan, 1994; Brun et al., 1994), we have adopted the partial equilibrium approach for the development of a reactive multi-component transport model (Prommer et al., 1999). In the present study, a multidimensional version of this model is applied in a qualitative sense for the interpretation of the ®eld observations described above, recognising that this approach does not allow us to model explicitly physiological controls (Lovley and Chapelle, 1995) on microbial metabolism. As a ®rst step, batch-type geochemical modelling
using PHREEQC (Parkhurst, 1995) was carried out for the background geochemistry of the ®eld site, with the aim of providing (equilibrated) initial conditions for the transport model. For this purpose, the background aqueous concentrations were taken from measurements at bore MP9 as shown in Table 2. The minerals assumed to be in equilibrium with the aqueous solution were siderite (FeCO3), pyrite (FeS2) and goethite (FeOOH). Goethite was assumed to represent the crystalline iron found in the sediment analysis following the results of qualitative XRD analysis of soil samples from within and below the `coee rock' layer (Whincup, 1994). Pyrite was assumed not to be present initially, as it was absent in the background core. Siderite was included because it was assumed to make up the largest part of the Fe(II) extracted by the 5 M HCl extraction. The initial mineral concentrations for goethite and siderite were taken as measured by the wet extractions. Given these assumptions, speciation modelling indicated that for the pH measured in the ®eld, Fe+2 and inorganic carbon concentrations were too low to stabilise siderite. As carbonate abundance was not quanti®ed directly during the acid digestion, a possible explanation for this might be that the Fe(II) extracted may have consisted of surface complexes or
H. Prommer et al. / Organic Geochemistry 30 (1999) 423±435
surface precipitates, which did not show up as exchangeable Fe(II) targeted by the 1 M CaCl extraction. Formation of such surface-associated Fe(II) has been noted, e.g. by Roden and Zachara (1996). Lovley and Phillips (1986b, 1987a) noted the formation of mixed Fe(III)-Fe(II) compounds, including magnetite, from the Fe(II) produced during reduction. Unfortunately, to our knowledge, no thermodynamic data for the formation of such compounds are currently available. Choosing magnetite (Fe3O4) as a representative of mixed Fe(III)-Fe(II) precipitates, for which thermodynamic data are available, equilibrium with goethite is possible. 6. Reactive transport modelling Following the results from the batch-type geochemical modelling above, we have developed three possible scenarios for the transport modelling study. For the ®rst scenario, S1, we assumed that the Fe(II) extracted was mainly siderite and we increased the initial total inorganic carbon concentration (TIC) and soluble Fe(II) concentrations until siderite was stable (Table 2). Thus, in the model, alkalinity and Fe(II) were overestimated compared to the concentrations measured in the ®eld. In the second scenario, S2, we assumed a groundwater composition as indicated in Table 2 with a modelled TIC, based on measured alkalinity and the pH measured in the ®eld. However, in this scenario, to maintain equilibrium, we have to assume that siderite was not present initially in the (uncontaminated) aquifer. Finally, a third case, S3, was modelled, in which goethite and magnetite were assumed to be present initially. Pyrite was allowed to precipitate, but assumed to be absent in the uncontaminated system. This case is thought to represent the case where non-speci®c, mixed Fe(II)-Fe(III) compounds have been produced during organic matter reduction before the aquifer was contaminated. To account to some extent for the additional Fe(III) fraction of magnetite, the initial concentration of goethite was reduced compared to S1 and S2. However, due to the Fe(III)/Fe(II) ratio of 1.6 within magnetite, the Fe(III) concentration is still overestimated compared to values measured by HCl extraction. 6.1. Model description The model applied for the above scenarios couples, via the operator-splitting method (Yeh and Tripathi, 1989; Barry et al., 1996; Steefel and MacQuarrie, 1996), three-dimensional advective-dispersive transport of multiple organic compounds and inorganic com-
431
ponents with the geochemical equilibrium package PHREEQC, a biodegradation module, and a dissolution module accounting for interphase mass-transfer from a non-aqueous phase liquid (NAPL) to an aqueous phase. Chemical reactions of inorganic species, including mineral dissolution/precipitation, are based on the local equilibrium assumption (LEA), which means that reactions are fast compared to hydrological transport. Redox equilibrium is achieved for inorganic species and minerals considered in the simulations but not for organic compounds (McNab and Narasimhan, 1994). The degradation reactions of organic compounds are assumed to be controlled by Monod-type kinetics and reaction rates depend on microbial activity. This is achieved by simulating microbial growth and decay of each microbial group and their subsequent uptake of organic substrates. The amount of organic compounds consumed during bacterial activity within a timestep Dt is assumed to be mineralised instantaneously and added to the inorganic system handled by PHREEQC. Thus, in the zones of ongoing degradation, oxidation capacity is consumed and inorganic carbon produced, thereby locally changing the redox state. 6.2. Model assumptions, initial and boundary conditions As indicated before, the groundwater ¯ow pattern of this particular ®eld site changes seasonally. A detailed simulation of these patterns was beyond the scope of this study. Thus, the simulation of the ¯ow-®eld underlying the transport problem was simpli®ed as a twodimensional steady state problem with an average groundwater ¯ow velocity of 135 m/yr (parallel ¯ow). Longitudinal and transversal dispersivities were 0.4 m and 0.04 m, respectively. The model area of 450 m (in ¯ow direction) 60 m was discretised by a regular grid with elements of 3 m (in ¯ow direction) 1 m in size. The chemical and physical properties of the aquifer were assumed to be homogeneous. The initial aqueous and mineral concentrations for the three modelling scenarios under consideration are listed in Table 2. The contamination source area (UST) was located towards the upstream boundary. The NAPL was distributed to 72 of the 9000 grid elements. The multicomponent NAPL mix consisted of six organic compounds (BTEX, propylbenzene, trimethylpentane), being dissolved by the passing uncontaminated groundwater. The dissolution process was assumed to be rate-limited and the multi-component solubility of the organic compounds was modelled according to Raoult's law (Hayden et al., 1992). For a better comparability of the three scenarios, only toluene was allowed to degrade in the model. It was most likely to have the greatest impact on the inorganic geochemistry, as: (i) rates of benzene degra-
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Fig. 5. Simulated aqueous concentrations of benzene, toluene, sulphate, Fe2+ and pH after 800 d (Scenario S3); the groundwater ¯ow direction is from the left to the right.
dation have been shown to be very low (Davis et al., 1995; Thierrin et al., 1995); and (ii) the dissolution rates of other more degradable compounds (e.g. ethylbenzene, xylenes) are signi®cantly lower due to their lower (multi-component) solubility. These compounds were treated as nonreactive tracers. 6.3. Results and discussion With the start of the simulation(s), organic compounds dissolve from the NAPL phase into the uncontaminated aquifer and form a steadily growing plume of dissolved organic compounds. After a short lag period, caused by the slow initial microbial growth, degradation/mineralisation rates of toluene increase and start to change signi®cantly the inorganic chemistry of the model aquifer. The degradation causes the toluene plumes to be retarded and to be narrower than the benzene plumes. This can be seen, e.g. for S3 in Fig. 5 for t=800 days after the start of the contamination. The model simulations reveal some signi®cant dierences between the modelling scenarios with respect to the Fe(III)/sulphate redox sequence in response to the mineral buering mechanisms. In the ®rst scenario (S1), as might be expected, iron reduction is thermodynamically more favourable and thus locally dominates over sulphate reduction where reducible iron is present. Due to the relatively low in-
itial goethite concentrations, the zone depleted in goethite grows rapidly, beginning near the contamination source. Sulphate reduction then takes over where goethite is exhausted and iron reduction becomes con®ned to the fringes of the toluene plume, mainly at the front of the plume. The greater the extent of the plume, the more important sulphate becomes as an electron acceptor, as hydrodynamic dispersion also mixes sulphate into the Fe(III)-depleted zone, providing additional oxidation capacity downstream. The Fe(III) reduced is largely converted to siderite, Fe2+ concentrations in the plume centre are slightly decreasing compared to initial concentrations. In this scenario, simulated pH and pe change only slightly in the contaminated zone. In the second scenario (S2), goethite is also reduced initially. However, as, in contrast to the ®rst scenario S1, the aqueous solution is not buered by siderite, the pH increases rapidly to about 5.5 at locations where toluene is degraded. At the same time, a switch to a preferential consumption of sulphate for any extra organic carbon added can be observed. Thus, in this scenario, goethite reduction is con®ned to a small region near the contamination source, while further downstream sulphate reduction is dominating. The sulphate reduced is converted to sulphide and precipitates as pyrite. Only very little siderite was produced in S2 and it was not formed at the initial stage of the simu-
H. Prommer et al. / Organic Geochemistry 30 (1999) 423±435
433
Fig. 6. Simulated concentrations of goethite, magnetite and pyrite after 800 d (Scenario S3); concentrations are expressed as mass per litre groundwater.
lation. It only forms where no sulphide is available to form pyrite. In the last scenario (S3), as in S1, goethite (locally) was reduced completely before the onset of sulphate reduction. With magnetite being the end product of iron reduction, there was no shift to sulphate reduction. However, as magnetite still contains 2/3 Fe(III) total electron ¯ow associated with iron reduction is smaller compared to S1. This scenario seems to provide a better analysis than S1 or S2, as it reproduces qualitatively the major geochemical patterns observed in the ®eld. Snapshots of selected organic compound and inorganic aqueous component, mineral and microbial concentrations are shown in Figs. 5, 6 and 7, respectively, for this case. 7. Conclusions Field investigations and model simulations were carried out in order to ®nd evidence as to whether Fe(III)-minerals contribute substantially to the oxi-
dation capacity of a shallow sand aquifer contaminated with petroleum hydrocarbons. The analysis of the sediments by wet extraction techniques showed that a large fraction of the sediments appeared to be in a reduced state before the onset of the contamination. The iron extractions revealed no clear evidence for a dierence in the Fe(III)/Fe(II) ratio between the uncontaminated and contaminated zone of the aquifer, while the pyrite found in the sediments con®rm ongoing sulphate reduction in the contaminated zone. Only low concentrations of reducible iron, mainly as 5 M HCl-extractable iron and thus a less reactive form (Roden and Zachara, 1996), contribute to the oxidation capacity of the aquifer. An electron balance calculation using the measured Fe(III) concentration has shown that the remaining Fe(III) had the potential to degrade the BTEX spill. However, geochemical modelling indicated that the iron found in the sediments could be mixed Fe(III)/Fe(II) compounds. These are thought not to be available for further reduction (Lovley and Phillips, 1986a). This was con®rmed by one of the transport modelling scenarios where magne-
Fig. 7. Simulated concentrations of Fe(III)-reducing and of sulphate-reducing bacteria after 800 d (Scenario S3); concentrations are expressed as mass per litre groundwater.
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tite was used as a representative of mixed Fe(III)/ Fe(II) compounds, the magnetite produced during goethite reduction was unavailable for further reduction. The transport model simulations have also shown that, under the geochemical conditions found at the ®eld site, the Fe(III)-sulphate redox sequence is, in addition to the form of the reducible Fe(III) mineral (Postma and Jakobsen, 1996), sensitive to the (assumed) end product of Fe(III) reduction, i.e. the Fe(II) minerals included in the simulation and their initial concentrations. The sensitivity of the modelling scenarios might also be used as an indicator that under natural conditions chemical heterogeneity, i.e. a suite of dierent Fe(II) and Fe(III) minerals, would lead to simultaneous occurrence of both iron and sulphate reduction. Clearly, due to the numerous idealisations made for both the batch-type and the reactive transport modelling, inferences from the modelling results have to be made carefully. In conclusion, iron appears not to play a signi®cant role as an electron acceptor at this particular ®eld site. The presence of iron still could have implications for the eciency of sulphate reduction since precipitation of sulphides prevents an increase of sulphide concentrations to toxic levels (Beller and Reinhard, 1995). It seems also, that the reduced forms of the electron acceptors become bound largely to the aquifer matrix by the formation of siderite, pyrite and/or magnetite. Thus the reduction capacity of the aquifer is further increased, with appropriate implications for aquifer remediation (Heron and Christensen, 1995; Heron et al., 1995). Acknowledgements This research was partly funded by the Centre for Groundwater Studies. We express our appreciation for the very helpful comments of the two anonymous reviewers. References Baedecker, M.J., Cozarelli, I.M., Siegel, D.I., Bennett, P.C., Eganhouse, R.P., 1993. Crude oil in a shallow sand and gravel aquifer. 3. Biogeochemical reactions and mass balance modeling. Appl. Geochem. 8 (6), 569±586. Barbaro, J.R., Barker, J.F., Lemon, L.A., May®eld, C.I., 1992. Biotransformation of BTEX under anaerobic, denitrifying conditions: Field and laboratory observations. J. Contam. Hydrol. 11, 245±272. Barber, C., Davis, G.B., Thierrin, J., Bates, L., Patterson, B.M., Pribac, F., Gibbs, R., Power, T.R., Briegel, D., Lambert, M., Hosking, J. 1991. Final report for project on `Assessment of the Impact of Pollutants on Groundwater beneath Urban Areas', July 1989 to June 1991. Technical report, CSIRO Division of Water Resources.
Barker, J.F., Patrick, G.C., Majo, D., 1987. Natural attenuation of aromatic hydrocarbons in a shallow sand aquifer. Ground Water Monitoring Review 7, 64±71. Barry, D.A., Bajracharya, K., Miller, C.T., 1996. Alternative split-operator approach for solving chemical reaction/ groundwater transport models. Adv. Water Resour. 19 (5), 261±275. Beller, H.R., Grbic-Galic, D., Reinhard, M., 1992. Microbial degradation of toluene under sulfate-reducing conditions and the in¯uence of iron on the process. Appl. Environ. Microbiol. 58 (16), 768±793. Beller, H.R., Reinhard, M., 1995. The role of iron in enhancing anaerobic toluene degradation in sulfate-reducing enrichment cultures. Microbial Ecology 30, 105±114. Brun, A., Engesgaard, P., Frind, E.O., 1994. A coupled microbiology-geochemistry transport model for saturated groundwater ¯ow. In: Dracos, T.H., Stauer, F. (Eds.), Transport and Reactive Processes in Aquifers. A.A. Balkema, pp. 457±462. Cline, J.D., 1969. Spectrophotometric determination of hydrogen sul®de in natural waters. Limnol. Oceanogr. 14, 454± 458. Davis, G.B., Barber, C., Power, T.R., Thierrin, J., Patterson, B.M., Rayner, J., Wu, Q. 1999. The variability and intrinsic remediation of a BTEX plume in anaerobic sulphaterich groundwater. J. Contam. Hydrol. 36 (3±4), 265±290. Davis, G.B., Power, T.R., Toze, S., Patterson, B.M., Johnston, C.D., Rayner, J., Whincup, T., McNaughton, N. 1995. Factors controlling the distribution of petroleum hydrocarbon plumes in groundwater at a site of gasoline leakage in Perth, Western Australia. Technical report, CSIRO Division of Water Resources. Edwards, E.A., Wills, L.E., Reinhard, M., Grbic-Galic, D., 1992. Anaerobic degradation of toluene and xylene by aquifer microorganisms under sulfate-reducing conditions. Appl. Environ. Microbiol. 58 (3), 794±800. Gerritse, R.G., Barber, C., Adeney, J.A. 1990. The impact of residential urban areas on groundwater quality: Swan Coastal Plain, Western Australia. Technical report, CSIRO, Water Resources Series. Grbic-Galic, D., Vogel, T.M., 1987. Transformation of toluene and benzene by mixed methanogenic cultures. Appl. Environ. Microbiol. 53, 254±260. Hayden, N., Voice, T.C., Annable, M.D., Wallace, R.B., 1992. Prediction of leachate concentrations in petroleumcontaminated soils. Journal of Soil Contamination 1, 81± 93. Heron, G., Bjerg, P.L., Christensen, T.H., 1995. Redox buffering in shallow aquifers contaminated by leachate. In: Hinchee, R.E., Wilson, J.T., Downey, D.C. (Eds.), Intrinsic bioremediation. Bioremediation, 3(1). Battelle Press, Columbus, OH, pp. 143±151. Heron, G., Christensen, T., 1995. Impact of sediment-bound iron on redox buering in a land®ll leachate polluted aquifer (Vejen, Denmark). Env. Sci. Technol. 29 (9), 1698± 1705. Heron, G., Crouzet, C., Bourg, A.C.M., Christensen, T., 1994. Speciation of Fe(II) and Fe(III) in contaminated aquifer sediments using chemical extraction techniques. Env. Sci. Technol. 28 (9), 1698±1705. Kuhn, E.P., Zeyer, J., Eicher, P., Schwarzenbach, R., 1986.
H. Prommer et al. / Organic Geochemistry 30 (1999) 423±435 Anaerobic degradation of alkylated benzene in denitrifying laboratory aquifer columns. Appl. Environ. Microbiol. 54, 490±496. Linge, K.L. 1996. Iron speciation in an aquifer contaminated by hydrocarbons. Department of Chemistry, University of Western Australia. Unpublished honours thesis. Lovley, D.R., Baedecker, M.J., Lonergan, D.J., Phillips, E.J.P., Siegel, D., 1989. Oxidation of aromatic contaminants coupled to microbial iron reduction. Nature 339, 297±299. Lovley, D.R., Chapelle, F.H., 1995. Deep subsurface microbial processes. Rev. Geophys. 33, 365±381. Lovley, D.R., Lonergan, D.J., 1990. Anaerobic oxidation of toluene, phenol, and p-cresol by the dissimilatory ironreducing organism, GS-15. Appl. Environ. Microbiol. 56 (6), 1858±1864. Lovley, D.R., Phillips, E.J.P., 1986a. Availability of ferric iron for microbial reduction in bottom sediments of the freshwater tidal Potomac river. Appl. Environ. Microbiol. 52 (4), 751±757. Lovley, D.R., Phillips, E.J.P., 1986b. Organic matter mineralization with reduction of ferric iron in anaerobic sediments. Appl. Environ. Microbiol. 51 (4), 683±689. Lovley, D.R., Phillips, E.J.P., 1987a. Anaerobic production of magnetite by a dissimilatory iron-reducing microorganism. Nature 330, 252±254. Lovley, D.R., Phillips, E.J.P., 1987b. Competitive mechanisms for inhibition of sulfate reduction and methane production in the zone of ferric iron reduction in sediments. Appl. Environ. Microbiol. 53 (11), 2636±2641. Ludvigsen, L., Heron, G., Albrechtsen, H.J., Christensen, T.H., 1995. Geomicrobial and geochemical redox processes in a land®ll-polluted aquifer. In: Hinchee, R.E., Wilson, J.T., Downey, D.C. (Eds.), Intrinsic bioremediation. Bioremediation, 3(1). Battelle Press, Columbus, OH, pp. 135±142. McNab, W.W., Narasimhan, T.N., 1994. Modeling reactive transport of organic compounds in groundwater using a partial redox disequilibrium approach. Water Resour. Res. 30 (9), 2619±2635. Parkhurst, D.L. 1995. User's guide to PHREEQCÐA computer program for speciation, reaction-path, advective-transport, and inverse geochemical calculations. Technical Report 4227, U.S. Geological Survey Water-Resources Investigations Report.
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Postma, D., Jakobsen, R., 1996. Redox Zonation: Equilibrium constraints on the Fe(III)/SO4Ðreduction interphase. Geochim. Cosmochim. Acta 60, 3169±3175. Prommer, H., Barry, D.A., Davis, G.B., 1999. A one-dimensional reactive multi-component transport model for biodegradation of petroleum hydrocarbons in groundwater. Env. Model. and Softw. 14 (2/3), 213±223. Roden, E., Zachara, J.M., 1996. Microbial reduction of crystalline iron(iii) oxides: In¯uence of oxide surface area and potential for cell growth. Env. Sci. Technol. 30 (9), 1698± 1705. Schmitt, R., Langguth, H.R., PuÈttmann, W., Rohns, H.P., Eckert, P., Schubert, J., 1996. Biodegradation of aromatic hydrocarbons under anoxic conditions in a shallow sand and gravel aquifer of the lower Rhine valley, Germany. Organic Geochemistry 25 (1/2), 41±50. Steefel, C.I., MacQuarrie, K.T.B., 1996. Approaches to modelling of reactive transport in porous media. In: Lichtner, P.C., Steefel, C.I., Oelkers, E.H. (Eds.), Reactive Transport in Porous Media: General Principles and Applications to Geochemical Processes. Mineralogical Society of America, pp. 83±129. Thierrin, J., Davis, G., Barber, C., 1995. A gound-water tracer test with deuterated compounds for monitoring in situ biodegradation and retardation of aromatic hydrocarbons. Ground Water 33 (3), 469±475. Whincup, T. 1994. Coee RockÐIts eect on the distribution and migration of a dissolved petroleum contaminant plume. Dept. of Geology, University of Western Australia. Unpublished honours thesis. Wisotzky, F., Eckert, P., 1997. Sulfat-dominierter BTEXAbbau im Grundwasser eines ehemalingen Gaswerksstandortes. Grundwasser 3 (1), 11±20. Yeh, G.T., Tripathi, V.S., 1989. A critical evaluation of recent developments in hydrogeochemical transport models of reactive multichemical components. Water Resour. Res. 25 (1), 93±108. Zehnder, A.J.B., Stumm, W., 1988. Geochemistry and biogeochemistry of anaerobic habitats. In: Zehnder, A.J.B. (Ed.), Biology of anaerobic microorganisms. Wiley, New York, pp. 1±38. Zeyer, J., Kuhn, E.P., Schwarzenbach, R., 1986. Rapid microbial mineralization of toluene and 1,3-dimethylbenzene in the absence of molecular oxygen. Appl. Environ. Microbiol. 52 (4), 944±947.
Geochemical changes during biodegradation of petroleum hydrocarbons: ®eld investigations and biogeochemical modelling H. Prommer a, b,*, G.B. Davis b, D.A. Barry c a
Department of Environmental Engineering, University of Western Australia, Nedlands, 6907, Australia b Centre for Groundwater Studies, CSIRO Land and Water, Wembley, 6014, Australia c School of Civil and Environmental Engineering and Contaminated Land Assessment and Remediation Research Centre, University of Edinburgh, Edinburgh, EH9 3JN, UK
Abstract Sediment analysis using wet extraction techniques and model simulations were carried out to investigate the role of ferric iron during biodegradation of dissolved petroleum hydrocarbon compounds at a ®eld site in Perth, Western Australia. Sediment cores were analysed for iron and sulphur species. The total iron concentrations were found to be low while the fraction of Fe(II) was surprisingly high. Pyrite was detected in the contaminated zone, but no iron monosulphides were found. A reactive multi-component transport model, coupling advective-dispersive transport of organic compounds and inorganic aqueous components with a geochemical equilibrium model and a biodegradation module, was applied to simulate qualitatively the ¯uxes and reactions involved in the biodegradation of BTEX compounds. Both ®eld investigations and modelling suggest that ferric iron minerals play no important role as electron acceptors while sulphate provides the major part of the oxidation capacity. # 1999 Elsevier Science Ltd. All rights reserved. Keywords: Groundwater; Iron reduction; Sulphate reduction; Multi-component transport; PHREEQC; BTEX
1. Introduction Petroleum products containing benzene, toluene, ethylbenzene and xylene (BTEX) have contaminated many shallow aquifers (Barker et al., 1987; Schmitt et al., 1996). Degradation of these compounds, catalysed by bacteria, has been shown to occur at dierent rates in many geochemical environments. BTEX degradation is well known in aerobic environments but is also reported to occur under nitrate-reducing (Zeyer et al.,
* Corresponding author. Fax: +61-8-9380-1015. E-mail address: [email protected] (H. Prommer)
1986; Kuhn et al., 1986; Barbaro et al., 1992), sulphate-reducing (Edwards et al., 1992; Beller et al., 1992; Wisotzky and Eckert, 1997), iron-reducing (Lovley et al., 1989; Lovley and Lonergan, 1990) and methane-producing (Grbic-Galic and Vogel, 1987) conditions. Generally, in conditions where neither the carbon-source nor nutrients limit microbial growth, the microbial communities consuming the thermodynamically most favourable electron acceptor outcompete other communities until the availability of the electron acceptor becomes limited. Thus, electron acceptors appear to be consumed sequentially, forming discrete redox-zones (Baedecker et al., 1993). However, in cases where both sulphate and Fe(III) serve as terminal elec-
0146-6380/99/$ - see front matter # 1999 Elsevier Science Ltd. All rights reserved. PII: S 0 1 4 6 - 6 3 8 0 ( 9 9 ) 0 0 0 2 7 - 3
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tron acceptors, the redox-sequence is not always clear. Lovley and Phillips (1987b) found inhibition of sulphate reduction by ongoing ferric iron reduction whereas Ludvigsen et al. (1995), for example, reported ongoing sulphate reduction in a land®ll rich in reducible iron sediments. Our study is aimed at identifying the role of ferric iron in the degradation process of BTEX at a particular contaminated shallow sand aquifer beneath Perth/Western Australia. More speci®cally we quantify the contribution of iron as a potential electron acceptor. If Fe(III) is not involved, then degradation at this ®eld site relies solely on sulphate as the electron acceptor implying that the observed degradation rates (Davis et al., 1999) are associated with sulphate reduction. 2. Field site The ®eld site is located in the Perth Metropolitan area. A leaking underground storage tank (UST) released approximately 500 l of gasoline, leading to a plume of more than 400 m in length for some compounds (see Figs. 1 and 2). Organic compounds and selected inorganic species have been monitored at the ®eld site since 1991 at up to 18 multiport boreholes with vertical spacings of 0.20±0.5 m (see Davis et al., 1999 for details). Biodegradation was shown to occur and (local) degradation rates were determined in situ via tracer tests using fully deuterated benzene, toluene, p-xylene and naphthalene (Thierrin et al., 1995). A detailed description of the contamination and its temporal and spatial changes is given by Davis et al. (1999). 2.1. Hydrogeology The site is part of the quaternary Bassendean Sand aquifer. At the ®eld site, the aquifer is 7 to 12 m thick and is underlain by a clay aquitard (Barber et al., 1991). The sands are generally medium-to-®ne grained bleached quartz sands. With the zone of water table ¯uctuation, organic matter has accumulated and the sands form a dark-brown, variably cemented 0.05±0.6m thick layer (Davis et al., 1999), locally called `coee rock'. The groundwater ¯ow direction changes seasonally by up to 208 while the water-table elevation varies by up to 1.8 m. Groundwater velocity at the ®eld site is estimated to be between 100 and 170 m/yr (Davis et al., 1999). 2.2. Local geochemistry The background geochemistry (see Table 1) of the contaminated site is generally anaerobic. Furthermore, nitrate concentrations are very low (0.1±1 mg/l,
Gerritse et al., 1990) though the residential area upstream of the study site is unsewered. Intensive denitri®cation is assumed to be the main reason for the low nitrate levels. In contrast to nitrate reduction, sulphate reduction is not thought to occur naturally in the Bassendean Sands (Gerritse et al., 1990). Thus, sulphate provides the largest potential oxidation capacity of the uncontaminated groundwater at the ®eld site. Typically, sulphate concentrations measured at MP9 (Fig. 1), a background multiport bore upstream of the underground storage tank, were between 50 and 100 mg/l. The average pH of background groundwater generally lies between 5 and 5.7. It decreases approximately 0.5 units between the top port and the bottom port of bore MP9, as can be seen in Fig. 2. Within the BTEX-contaminated zone, sulphate concentrations were signi®cantly lower than outside the plume. This can be seen on the contour plot for August 1994 (Fig. 1), in the appropriate depth pro®les along the main ¯owline (Fig. 2) and in Table 1, where mean values of aqueous concentrations (organic and inorganic compounds) outside and within the plume are listed for comparison. The lowest sulphate concentrations were generally found close to the contamination source (MP12, MP11). Hydrogen sulphide was found in low concentrations (<3 mg/l) in samples from within the contaminant plume. Bicarbonate concentrations were signi®cantly higher within the plume whereas no clear trend was exhibited for Fe(II). Despite the minor dierence in mean values (Table 1), signi®cantly increased Fe(II) concentrations (compared to background values) were occasionally (especially in the ®rst few years after the BTEX spill) measured in some ports of wells closer to the contamination source. In contrast, Fe(II) concentrations further downstream, e.g. at MP6, seemed to be lower within the contaminated zone (Fig. 2). The pH increased typically by 0.2±0.7 units within the plume while measured Eh values, as expected, decreased (Table 1). 3. Sediment analysis 3.1. Methodology 3.1.1. Sampling For this study, two cores were recovered from the ®eld site in November 1996. One was taken from the uncontaminated zone up-gradient of the spill halfway between the spill (UST on Fig. 1) and bore MP9. It was thought to provide background concentrations. The other core was recovered from the contaminated zone, 80 m downstream from the spill area next to bore MP1. As the BTEX plume is relatively thin, it was expected that the core from the contaminated
Fig. 1. Benzene, toluene and sulphate concentrations at the ®eld site in August 1994 (Davis et al., 1995),
H. Prommer et al. / Organic Geochemistry 30 (1999) 423±435 425
Fig. 2. Depth pro®les of selected organic compounds and inorganic species in August 1994; Depth=depth below water table.
426 H. Prommer et al. / Organic Geochemistry 30 (1999) 423±435
H. Prommer et al. / Organic Geochemistry 30 (1999) 423±435 Table 1 Mean values for inorganic compounds and physical parameters inside and outside of the contaminant plume (Linge, 1996) Parameter
Meanbackgr.a
Meaninsideb
O2 (mg/l) NOÿ 3 (mg/l) NH+ 4 (mg/l) Fe2+ (mg/l) SO2ÿ (mg/l) 4 H2S (mg/l) HCOÿ 3 (mg/l) pH Eh (mV)
0.56 0.18 0.25 1.6 82.5 0.1 14.3 5.1 103
0.16 0.05 0.3 1.4 32.4 1.1 74 5.7 ÿ 30
a
Total BTEX concentration R1 mg/l, 102 samples, except + for NOÿ 3 and NH4 (30 samples) and H2S (20 samples). b Total BTEX concentration >1 mg/l, 170 samples, except + for NOÿ 3 and NH4 (80 samples) and H2S (54 samples).
zone would show a pronounced depletion of reducible iron in the vertical pro®le. The cores were recovered in aluminium tubing (5-cm diameter) using a hollow-stem auger drilling rig. The aluminium tubing was cut into 25-cm sections, sealed immediately and placed within an anaerobic chamber within several hours. At the time of sampling, the groundwater table of both cores was located between the top section (25-cm section) and the second section. The bottom sections of both cores are taken from a depth of approximately 2.5 m below the appropriate groundwater table. The top 10 cm and bottom 10 cm of each section was discarded to minimise the in¯uence of any reoxidation taking place at the core ends. The remaining sediment material was then mixed thoroughly and subsamples of 2±3 g were placed in 20-ml glass serum bottles. 3.1.2. Analytical procedure The sediments were analysed and concentrations quanti®ed using the wet extraction techniques as tested and described by Heron et al. (1994). Extractions using 1 M CaCl, 0.5 M HCl and 5 M HCl were carried out, corresponding to ion-exchangeable, amorphous and crystalline forms of iron, respectively. After addition of the appropriate extractant (10 ml) in the anaerobic chamber, the glass bottles were sealed with a rubber stopper. Samples tested for ion-exchangeable iron were rotated for 8 h before being analysed for Fe(II) and Fe(III). The extraction time of the 0.5 M HCl extraction was 24 h. The 5 M HCl-extracted samples were heated for 1 h at 1208C. This procedure (Linge, 1996) modi®es slightly that of Heron et al. (1994) and was found to dissolve iron more completely. In particular, the method was able to completely extract even Fe(II) carbonates. For each sampling depth, triplicates were
427
analysed for Fe(II) and the total iron concentration (Fetot) by a colorimetric method using ferrozine; absorbance was determined with a VARIAN 100 DMS spectrophotometer. Quanti®cation of total iron concentrations was carried out by reducing iron completely to Fe(II) with hydroxylamine hydrochloride. Fe(III) concentrations were obtained from the dierence between Fetot and Fe(II). AVS (acid volatile sulphur) was analysed to determine FeS concentrations (Heron et al., 1994), assuming that AVS was all FeS. A sequential extraction by HI, and a 0.7 M Cr(II), 2 M HCl mixture was used for FeS2. Extractions were carried out using a digestion apparatus in which the H2S released was trapped in a zinc acetate solution. The methylene blue method (Cline, 1969) was used to quantify the extracted sulphides. Pyrite concentrations were then calculated from the dierence between HI-extracted sulphides and the Cr(II)-extracted sulphides. 3.2. Results The methodology described above resolved the depth pro®les of iron in the two cores taken. The errors within the triplicate samples typically were smaller than the dierences between two subsequent sampling depths. However, concentrations of ionexchangeable, amorphous and crystalline iron in both cores were lower than initially expected. A maximum total iron concentration of 900 mgFe/gsoil was found by 5 M HCl extraction for the yellow sands at the top of the background core (Fig. 3), recovered outside the BTEX plume. Below the yellow sand, two samples between 17 and 17.5 m AHD showed a bleached, white sand with minimal iron concentrations. Below these white sands, two samples showed dark brown `coee rock' with concentrations above 200 mgFe/gsoil of 5 M HCl-extractable iron. Concentrations less than 200 mgFe/gsoil were found in samples of light brown sand below the `coee rock'. In the core taken from the contaminated zone, the top four subsamples (16.5 and 17.5 m AHD) consisted of bleached white sand followed by one subsample of dark brown `coee rock' (approx. 16.25 m AHD), having the maximum total iron concentration of that core (200 mgFe/gsoil). Below that depth, total iron concentrations dropped back steadily as the sand became more and more bleached. From the total iron concentrations found, the largest fraction seemed to be Fe(II), even in the background core. This indicates that a large part of the reactive Fe(III) oxides were reduced before the aquifer was contaminated by BTEX, presumably by oxidation of natural organic matter. Reducible iron was mainly found as 5 M HCl-extractable iron in the background core, suggesting that Fe(III) minerals mainly occur in crystalline form. The concentrations were almost all
428
H. Prommer et al. / Organic Geochemistry 30 (1999) 423±435
Fig. 3. Concentrations of 5 M HCl and of 0.5 M HCl-extractable iron in units of m/g of soil; values for Fe(III) were calculated from Fetot ÿ Fe(II); AHD=Australian Height Datum.
around 40±60 mgFe/gsoil (see Fig. 3). Ion-exchangeable iron concentrations (not shown) were insigni®cant in all samples. Iron monosulphide concentrations (FeS, concentrations not shown) were not signi®cant in either core. Pyrite was present in 5 of 10 samples in the core taken
from the contaminated zone, but in none of the samples from the background core. However, the sequential HI/Cr(II) attack showed a high variability amongst the triplicates analysed, especially compared to the other extraction methods. In all background core samples and in some locations in the core from
H. Prommer et al. / Organic Geochemistry 30 (1999) 423±435
the contaminated area, HI-extracted sulphide concentrations were higher than the concentrations extracted by Cr(II) attack (Fig. 4).
4. Electron balance An electron balance calculation was carried out to verify whether the amount of Fe(III) found in the analysis would be sucient to account for the observed degradation of the BTEX compounds. We therefore assume that all Fe(III), including the fraction extracted by 5 M HCl, is available for the oxidation of the organic contaminants. This assumption is supported by observations of aquifer sediments contaminated by leachate from a Danish land®ll (Heron and Christensen, 1995). There, it was suggested that, over time, most of the crystalline Fe(III) oxide content can be utilised for organic matter degradation. In addition, Roden and Zachara (1996) noted that natural crystalline oxides may function as more eective electron acceptors than has been recognised previously. For the electron balance, we consider only electrondonating reactions of BTEX compounds, e.g. for toluene: ÿ C7 H8 21H2 O47COÿ2 3 50H 36e ,
1
while, for the electron-accepting reaction, we consider reduction of goethite: FeOOH eÿ 3H 4Fe2 2H2 O:
429
lents/m3) were available for the BTEX degradation, a bulk volume of roughly 44,300 m3 would be needed to achieve complete mineralisation of the BTEX compounds by iron reduction only. Considering the seasonal variability of the groundwater ¯ow direction, this is in fact approximately the bulk volume in contact with the BTEX plume. A rough estimate for an annual electron ¯ow, Eas, due to sulphate reduction, ÿ ÿ SOÿ2 4 9H 8e 4HS 4H2 O,
3
neglecting dispersive ¯uxes, might be derived from: Eas vne Ap DCSOÿ2 nt , 4
4
where v is the pore water velocity, ne is the (eective) porosity, Ap a cross-sectional area, nt is the number of electrons transferred and DCSOÿ2 is the average change 4 in sulphate concentration due to sulphate reduction within water passing through the cross-sectional area Ap. With groundwater ¯ow velocities of 100±170 m/yr, ne and nt of 0.3 and 8, respectively, and estimates for Ap and DCSOÿ2 of 60 m2 and 20 mg/l respectively, 4 between 4000 and 6700 equivalents are consumed per year. The latter estimate indicates that iron-reduction certainly cannot be the only electron-accepting process. However, it becomes also clear that, despite the low concentrations of Fe(III) oxides, there was potential for a signi®cant contribution of iron as an electron acceptor, if it could locally outcompete sulphate reduction.
2
Neglecting any biodegradation, vaporisation or other processes reducing the total contaminant mass while it passes through the unsaturated zone, we might estimate that 1600 mol of BTEX or approximately 57,600 equivalents (using toluene as a model compound) have reached the saturated zone. Assuming that 40 mgFe(III)/gsoil (=1.3 mol of Fe(III)/m3 or 1.3 equiva-
5. Geochemical equilibrium considerations As pointed out by Postma and Jakobsen (1996), microbially mediated degradation reactions can be seen as a two-step process, a ®rst step being the fermentation of, e.g. organic matter and the second being a subsequent consumption of the fermentation products
Fig. 4. Concentrations of sulphur in units of mg/g of soil (lines); Cr(II) extraction=D, HI=w; AHD=Australian Height Datum.
430
H. Prommer et al. / Organic Geochemistry 30 (1999) 423±435
Table 2 Measured and equilibrated aqueous and mineral concentrations Aqueous component,a Mineral
Cbackgr. measured (mol/l)b
Cbackgr., Cin¯ow equilibrated (mol/l)b, (S1)
Cbackgr., Cin¯ow equilibrated (mol/l)b, (S2)
Cbackgr., Cin¯ow equilibrated (mol/l)b, (S3)
pH pc Eh (mV) C(IV) Alkalinity S(VI) S(ÿII) Fe(II) Fe(III) Ca Mg Na K Cl Fe(III) Fe(II) Goethitec Siderite Pyrite Magnetite
5.0 ± 129
5.21 ÿ 0.45 ± 3.69 10ÿ2 2.97 10ÿ3 7.84 10ÿ4 5.27 10ÿ12 1.33 10ÿ3 1.61 10ÿ12 8.41 10ÿ4 5.66 10ÿ4 5.24 10ÿ3 2.59 10ÿ4 6.45 10ÿ3
5.07 1.21 ± 7.95 10ÿ3 4.29 10ÿ4 7.84 10ÿ4 9.27 10ÿ24 6.65 10ÿ5 2.14 10ÿ12 8.41 10ÿ4 5.66 10ÿ4 5.24 10ÿ3 2.59 10ÿ4 6.45 10ÿ3
5.07 1.21 ± 7.95 10ÿ3 4.29 10ÿ4 7.84 10ÿ4 9.27 10ÿ24 6.65 10ÿ5 2.14 10ÿ12 8.41 10ÿ4 5.66 10ÿ4 5.24 10ÿ3 2.59 10ÿ4 6.45 10ÿ3
1.00 10ÿ3 4.00 10ÿ3 0.0 0.0
1.00 10ÿ3 0.0 0.0 0.0
1.00 10ÿ3 ± 0.0 4.00 10ÿ3
3.27 10ÿ4 7.83 10ÿ4 < 3.0 10ÿ6 6.65 10ÿ5 ± 8.41 10ÿ4 5.65 10ÿ4 5.23 10ÿ3 2.58 10ÿ4 6.44 10ÿ3 0.4±1.2 10ÿ3 2.0±4.0 10ÿ3 0.0
a
Values in parentheses indicate valence. If not indicated otherwise. c Representing crystalline iron.
b
(e.g. acetate, formate, H2) by the dierent terminal electron-accepting processes (TEAPs). Postma and Jakobsen (1996) also argue that, if fermentation is the rate-controlling step in the overall degradation reaction, the reactions can be described as a partial equilibrium process (Zehnder and Stumm, 1988) in which the electron-accepting process is an equilibrium reaction. They have used that approach for geochemical equilibrium considerations at the Fe(III)/sulfate-reduction interface and show that, in dierent geochemical environments (depending on the stability of the Fe(III) oxides and the pH), either sulphate reduction or iron reduction could be thermodynamically more favourable. As in other models (McNab and Narasimhan, 1994; Brun et al., 1994), we have adopted the partial equilibrium approach for the development of a reactive multi-component transport model (Prommer et al., 1999). In the present study, a multidimensional version of this model is applied in a qualitative sense for the interpretation of the ®eld observations described above, recognising that this approach does not allow us to model explicitly physiological controls (Lovley and Chapelle, 1995) on microbial metabolism. As a ®rst step, batch-type geochemical modelling
using PHREEQC (Parkhurst, 1995) was carried out for the background geochemistry of the ®eld site, with the aim of providing (equilibrated) initial conditions for the transport model. For this purpose, the background aqueous concentrations were taken from measurements at bore MP9 as shown in Table 2. The minerals assumed to be in equilibrium with the aqueous solution were siderite (FeCO3), pyrite (FeS2) and goethite (FeOOH). Goethite was assumed to represent the crystalline iron found in the sediment analysis following the results of qualitative XRD analysis of soil samples from within and below the `coee rock' layer (Whincup, 1994). Pyrite was assumed not to be present initially, as it was absent in the background core. Siderite was included because it was assumed to make up the largest part of the Fe(II) extracted by the 5 M HCl extraction. The initial mineral concentrations for goethite and siderite were taken as measured by the wet extractions. Given these assumptions, speciation modelling indicated that for the pH measured in the ®eld, Fe+2 and inorganic carbon concentrations were too low to stabilise siderite. As carbonate abundance was not quanti®ed directly during the acid digestion, a possible explanation for this might be that the Fe(II) extracted may have consisted of surface complexes or
H. Prommer et al. / Organic Geochemistry 30 (1999) 423±435
surface precipitates, which did not show up as exchangeable Fe(II) targeted by the 1 M CaCl extraction. Formation of such surface-associated Fe(II) has been noted, e.g. by Roden and Zachara (1996). Lovley and Phillips (1986b, 1987a) noted the formation of mixed Fe(III)-Fe(II) compounds, including magnetite, from the Fe(II) produced during reduction. Unfortunately, to our knowledge, no thermodynamic data for the formation of such compounds are currently available. Choosing magnetite (Fe3O4) as a representative of mixed Fe(III)-Fe(II) precipitates, for which thermodynamic data are available, equilibrium with goethite is possible. 6. Reactive transport modelling Following the results from the batch-type geochemical modelling above, we have developed three possible scenarios for the transport modelling study. For the ®rst scenario, S1, we assumed that the Fe(II) extracted was mainly siderite and we increased the initial total inorganic carbon concentration (TIC) and soluble Fe(II) concentrations until siderite was stable (Table 2). Thus, in the model, alkalinity and Fe(II) were overestimated compared to the concentrations measured in the ®eld. In the second scenario, S2, we assumed a groundwater composition as indicated in Table 2 with a modelled TIC, based on measured alkalinity and the pH measured in the ®eld. However, in this scenario, to maintain equilibrium, we have to assume that siderite was not present initially in the (uncontaminated) aquifer. Finally, a third case, S3, was modelled, in which goethite and magnetite were assumed to be present initially. Pyrite was allowed to precipitate, but assumed to be absent in the uncontaminated system. This case is thought to represent the case where non-speci®c, mixed Fe(II)-Fe(III) compounds have been produced during organic matter reduction before the aquifer was contaminated. To account to some extent for the additional Fe(III) fraction of magnetite, the initial concentration of goethite was reduced compared to S1 and S2. However, due to the Fe(III)/Fe(II) ratio of 1.6 within magnetite, the Fe(III) concentration is still overestimated compared to values measured by HCl extraction. 6.1. Model description The model applied for the above scenarios couples, via the operator-splitting method (Yeh and Tripathi, 1989; Barry et al., 1996; Steefel and MacQuarrie, 1996), three-dimensional advective-dispersive transport of multiple organic compounds and inorganic com-
431
ponents with the geochemical equilibrium package PHREEQC, a biodegradation module, and a dissolution module accounting for interphase mass-transfer from a non-aqueous phase liquid (NAPL) to an aqueous phase. Chemical reactions of inorganic species, including mineral dissolution/precipitation, are based on the local equilibrium assumption (LEA), which means that reactions are fast compared to hydrological transport. Redox equilibrium is achieved for inorganic species and minerals considered in the simulations but not for organic compounds (McNab and Narasimhan, 1994). The degradation reactions of organic compounds are assumed to be controlled by Monod-type kinetics and reaction rates depend on microbial activity. This is achieved by simulating microbial growth and decay of each microbial group and their subsequent uptake of organic substrates. The amount of organic compounds consumed during bacterial activity within a timestep Dt is assumed to be mineralised instantaneously and added to the inorganic system handled by PHREEQC. Thus, in the zones of ongoing degradation, oxidation capacity is consumed and inorganic carbon produced, thereby locally changing the redox state. 6.2. Model assumptions, initial and boundary conditions As indicated before, the groundwater ¯ow pattern of this particular ®eld site changes seasonally. A detailed simulation of these patterns was beyond the scope of this study. Thus, the simulation of the ¯ow-®eld underlying the transport problem was simpli®ed as a twodimensional steady state problem with an average groundwater ¯ow velocity of 135 m/yr (parallel ¯ow). Longitudinal and transversal dispersivities were 0.4 m and 0.04 m, respectively. The model area of 450 m (in ¯ow direction) 60 m was discretised by a regular grid with elements of 3 m (in ¯ow direction) 1 m in size. The chemical and physical properties of the aquifer were assumed to be homogeneous. The initial aqueous and mineral concentrations for the three modelling scenarios under consideration are listed in Table 2. The contamination source area (UST) was located towards the upstream boundary. The NAPL was distributed to 72 of the 9000 grid elements. The multicomponent NAPL mix consisted of six organic compounds (BTEX, propylbenzene, trimethylpentane), being dissolved by the passing uncontaminated groundwater. The dissolution process was assumed to be rate-limited and the multi-component solubility of the organic compounds was modelled according to Raoult's law (Hayden et al., 1992). For a better comparability of the three scenarios, only toluene was allowed to degrade in the model. It was most likely to have the greatest impact on the inorganic geochemistry, as: (i) rates of benzene degra-
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Fig. 5. Simulated aqueous concentrations of benzene, toluene, sulphate, Fe2+ and pH after 800 d (Scenario S3); the groundwater ¯ow direction is from the left to the right.
dation have been shown to be very low (Davis et al., 1995; Thierrin et al., 1995); and (ii) the dissolution rates of other more degradable compounds (e.g. ethylbenzene, xylenes) are signi®cantly lower due to their lower (multi-component) solubility. These compounds were treated as nonreactive tracers. 6.3. Results and discussion With the start of the simulation(s), organic compounds dissolve from the NAPL phase into the uncontaminated aquifer and form a steadily growing plume of dissolved organic compounds. After a short lag period, caused by the slow initial microbial growth, degradation/mineralisation rates of toluene increase and start to change signi®cantly the inorganic chemistry of the model aquifer. The degradation causes the toluene plumes to be retarded and to be narrower than the benzene plumes. This can be seen, e.g. for S3 in Fig. 5 for t=800 days after the start of the contamination. The model simulations reveal some signi®cant dierences between the modelling scenarios with respect to the Fe(III)/sulphate redox sequence in response to the mineral buering mechanisms. In the ®rst scenario (S1), as might be expected, iron reduction is thermodynamically more favourable and thus locally dominates over sulphate reduction where reducible iron is present. Due to the relatively low in-
itial goethite concentrations, the zone depleted in goethite grows rapidly, beginning near the contamination source. Sulphate reduction then takes over where goethite is exhausted and iron reduction becomes con®ned to the fringes of the toluene plume, mainly at the front of the plume. The greater the extent of the plume, the more important sulphate becomes as an electron acceptor, as hydrodynamic dispersion also mixes sulphate into the Fe(III)-depleted zone, providing additional oxidation capacity downstream. The Fe(III) reduced is largely converted to siderite, Fe2+ concentrations in the plume centre are slightly decreasing compared to initial concentrations. In this scenario, simulated pH and pe change only slightly in the contaminated zone. In the second scenario (S2), goethite is also reduced initially. However, as, in contrast to the ®rst scenario S1, the aqueous solution is not buered by siderite, the pH increases rapidly to about 5.5 at locations where toluene is degraded. At the same time, a switch to a preferential consumption of sulphate for any extra organic carbon added can be observed. Thus, in this scenario, goethite reduction is con®ned to a small region near the contamination source, while further downstream sulphate reduction is dominating. The sulphate reduced is converted to sulphide and precipitates as pyrite. Only very little siderite was produced in S2 and it was not formed at the initial stage of the simu-
H. Prommer et al. / Organic Geochemistry 30 (1999) 423±435
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Fig. 6. Simulated concentrations of goethite, magnetite and pyrite after 800 d (Scenario S3); concentrations are expressed as mass per litre groundwater.
lation. It only forms where no sulphide is available to form pyrite. In the last scenario (S3), as in S1, goethite (locally) was reduced completely before the onset of sulphate reduction. With magnetite being the end product of iron reduction, there was no shift to sulphate reduction. However, as magnetite still contains 2/3 Fe(III) total electron ¯ow associated with iron reduction is smaller compared to S1. This scenario seems to provide a better analysis than S1 or S2, as it reproduces qualitatively the major geochemical patterns observed in the ®eld. Snapshots of selected organic compound and inorganic aqueous component, mineral and microbial concentrations are shown in Figs. 5, 6 and 7, respectively, for this case. 7. Conclusions Field investigations and model simulations were carried out in order to ®nd evidence as to whether Fe(III)-minerals contribute substantially to the oxi-
dation capacity of a shallow sand aquifer contaminated with petroleum hydrocarbons. The analysis of the sediments by wet extraction techniques showed that a large fraction of the sediments appeared to be in a reduced state before the onset of the contamination. The iron extractions revealed no clear evidence for a dierence in the Fe(III)/Fe(II) ratio between the uncontaminated and contaminated zone of the aquifer, while the pyrite found in the sediments con®rm ongoing sulphate reduction in the contaminated zone. Only low concentrations of reducible iron, mainly as 5 M HCl-extractable iron and thus a less reactive form (Roden and Zachara, 1996), contribute to the oxidation capacity of the aquifer. An electron balance calculation using the measured Fe(III) concentration has shown that the remaining Fe(III) had the potential to degrade the BTEX spill. However, geochemical modelling indicated that the iron found in the sediments could be mixed Fe(III)/Fe(II) compounds. These are thought not to be available for further reduction (Lovley and Phillips, 1986a). This was con®rmed by one of the transport modelling scenarios where magne-
Fig. 7. Simulated concentrations of Fe(III)-reducing and of sulphate-reducing bacteria after 800 d (Scenario S3); concentrations are expressed as mass per litre groundwater.
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tite was used as a representative of mixed Fe(III)/ Fe(II) compounds, the magnetite produced during goethite reduction was unavailable for further reduction. The transport model simulations have also shown that, under the geochemical conditions found at the ®eld site, the Fe(III)-sulphate redox sequence is, in addition to the form of the reducible Fe(III) mineral (Postma and Jakobsen, 1996), sensitive to the (assumed) end product of Fe(III) reduction, i.e. the Fe(II) minerals included in the simulation and their initial concentrations. The sensitivity of the modelling scenarios might also be used as an indicator that under natural conditions chemical heterogeneity, i.e. a suite of dierent Fe(II) and Fe(III) minerals, would lead to simultaneous occurrence of both iron and sulphate reduction. Clearly, due to the numerous idealisations made for both the batch-type and the reactive transport modelling, inferences from the modelling results have to be made carefully. In conclusion, iron appears not to play a signi®cant role as an electron acceptor at this particular ®eld site. The presence of iron still could have implications for the eciency of sulphate reduction since precipitation of sulphides prevents an increase of sulphide concentrations to toxic levels (Beller and Reinhard, 1995). It seems also, that the reduced forms of the electron acceptors become bound largely to the aquifer matrix by the formation of siderite, pyrite and/or magnetite. Thus the reduction capacity of the aquifer is further increased, with appropriate implications for aquifer remediation (Heron and Christensen, 1995; Heron et al., 1995). Acknowledgements This research was partly funded by the Centre for Groundwater Studies. We express our appreciation for the very helpful comments of the two anonymous reviewers. References Baedecker, M.J., Cozarelli, I.M., Siegel, D.I., Bennett, P.C., Eganhouse, R.P., 1993. Crude oil in a shallow sand and gravel aquifer. 3. Biogeochemical reactions and mass balance modeling. Appl. Geochem. 8 (6), 569±586. Barbaro, J.R., Barker, J.F., Lemon, L.A., May®eld, C.I., 1992. Biotransformation of BTEX under anaerobic, denitrifying conditions: Field and laboratory observations. J. Contam. Hydrol. 11, 245±272. Barber, C., Davis, G.B., Thierrin, J., Bates, L., Patterson, B.M., Pribac, F., Gibbs, R., Power, T.R., Briegel, D., Lambert, M., Hosking, J. 1991. Final report for project on `Assessment of the Impact of Pollutants on Groundwater beneath Urban Areas', July 1989 to June 1991. Technical report, CSIRO Division of Water Resources.
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