The additional groundwater quality data collected over the past three years has continued to establish trends and provide a broad basis for more detailed geochemical analysis. The continuing trends for SO4, Cl and Na are presented in Figure 3.12, and those for TDS and pH in Figure 3.13.
The trends of groundwater quality illustrate that SO4 at bore 2124U is rising at a similar rate to the RWP, but SO4 is yet to be detected in bore 3135U. The increase in SO4 corresponds to similar increases in TDS, showing the signature of ash pond water chemistry. The data collected over the last three years has shown some variation in SO4 at bore 2124U and the RWP, resembling the variations in SO4 in LYAP water. Importantly, the SO4 at the RWP is now approaching that in the LYAP. This suggests that the attenuation capacity of the HHF aquifer sediments, principally controlled by organic content, may be becoming limited due to consumption of the more reactive organic species in the vicinity of the RWP and beneath the LYAP.
Figure 3.13 - Trends of TDS and pH to January 1999
The Cl concentration of the RWP and bore 2124U continues to be stable, approximately at the same concentration as the LYAP with a small upward trend apparent. The Cl concentration in bore 3135U began to decrease and now appears relatively stable. The Cl concentration of bore 3135U is still similar to that of 2124U and the RWP. This may reflect the slow increase in Cl within the LYAP. The assumption of conservative migration of Cl is therefore still considered to be valid.
The CO2 (aq) content of bore 3135U is much higher than attributable to SO4 reduction alone, since the concentration of SO4 reaching this bore should be low according to the transport analysis in Section 3.5.6. Based on equation 3-11, the molar ratio of CO2 (or HCO3-) to SO4 is 2:1. The molar concentration of SO4 is about 0.01 mmol/L (using the 1 mg/L detection limit), compared to the molar concentration of CO2 at about 44 mmol/L. Even if it is assumed that the maximum groundwater velocity gives a SO4 concentration of about 1,000 mg/L, this is still only about 10.4 mmol/L. It is difficult therefore to explain the origin of this extra CO2. It is possible that it is related to degradation of reactive organic species in HHF aquifer sediments by competing bacteria (Knox et al., 1993; Manahan, 1991). The conversion of N and P may indicate further bacterial activity in HHF sediments, although this hypothesis is untested. The recent field data obtained does not allow any interpretation or assessment of possible mechanisms or processes.
The Na concentration is steadily increasing at the RWP, but remains at about background concentrations in bores 2124U and 3135U. The clay content of the HHF sediments is dominantly kaolinite, with minor illite and montmorillonite (Bolger, 1984). The retardation of Na migration is likely to be due to ion exchange and soprtion on these clay surfaces (cf. Fetter, 1993; McBride, 1994; Langmuir, 1997). Without further field or laboratory data it is not possible to assess the cation exchange capacity (CEC) of the HHF aquifer sediments and thus the potential for sodium sorption. The study by Mulder & Pedler (1990) adopted a retardation coefficient of 1.1, based on visual calibration to monitoring data. Their study predicted Na concentrations by 1995 of 670 mg/L in the 2124U area - markedly higher than the 240 to 290 mg/L detected around this time. Since they adopted the same retardation coefficient for Na as that used for SO4, the only difference in applying the analytical solute transport model from Section 3.5.6 is the source concentration from LYAP seepage to the HHF. The field data for Na is included in Figure 3.11, and shows a poor correlation to the analytical model and properites adopted. Further study of Na behaviour is necessary to properly quantify and assess the migration of Na in HHF aquifers.
The average oxidisable organic carbon (OOC) of HHF aquifer sediments has been determined to be about 2.5%. The source of this organic matter is expected to be reworked coal and peat material from the underlying coal seams during fluvial deposition of the HHF sediments. This provides a bountiful source of organic carbon for SO4-reducing bacteria. The reactive proportion of this organic material available to SRB, however, is still unknown. Given that peat material is likely to be of lower reactivity, as discussed in Section 3.6.2, it is expected that only a minor proportion of this 2.5% is utilised by the SRB. This correlates with the continuing increase in SO4 at the RWP and the possibility that the more reactive organic fraction has now been consumed beneath the LYAP and is providing less attenuation of SO4 as a result. A more detailed study is recommended that can identify the chief organic species present in HHF sediments and their respective geochemical reactivity rates with regards to facilitating SO4 reduction.
One approach which can be used in the interim is the application of a solute transport model which incorporates the reactive transport of SO4. This can be used to estimate average reaction rates by calibration to existing field data, and is a popular approach in the literature for a range of marine and sedimentary environments (Berner, 1964, 1980, 1981; Jørgensen, 1978c; Westrich & Berner, 1984; Berner & Westrich, 1985; Middelburg, 1989; Boudreau & Ruddick, 1991). The LYAP has nearly 20 years of monitoring data and thus a solute transport model could be applied by incorporating the kinetic reactions for SO4 reduction. This approach will be developed for groundwater flow and applied to the western seepage pathway later in this chapter, since most of the cited literature above refers to marine or estuarine sediments.
The nutrient concentrations appear to be sufficient to sustain the activity of SO4- reducing bacteria (cf. McNab & Dunlap, 1975; Knox et al., 1993; Bedient et al., 1994). Importantly, the chemical speciation undertaken on N and P shows that they are generally in forms of reduced NH3 and P. This may be explained through a more idealised representation of organic matter decomposition during SO4 reduction which includes N and P in the organic fraction. Knox et al. (1993) present the average composition of organic biomass as C60H87O23N12P.
For SO4 reduction, this formula for organic matter is further simplified and represented as (Richards, 1965; Murray, 1978; Boudreau & Westrich, 1984) :
2(CH2O)x(NH3)y(H3PO4)z + xSO42-→ 2xHCO3- + xH2S + 2yNH3 + 2zH3PO4 3-12 where x, y and z are constants, depending on the microorganisms' specific requirements.
This might help to explain the presence of N as NH3, however, it does not account for the speciation of P. It is possible for some of this NH3 to be oxidised to form NO3, given the HHF aquifer has low concentrations of oxygen present. The presence of NO3 was observed in the monitoring data collected. The reaction, known as nitrification, is generally given as (Manahan, 1991) :
2O2 + NH4+→ NO3- + 2H+ + H2O 3-13
The process of nitrification is also mediated by bacteria, with NO2 being produced as an intermediate metabolic by-product (Manahan, 1991). The presence of NO2 was generally not detected in most groundwater samples, except for bore 2173U in April 1998 (see Table 3.10). Bore 2173U also had one of the highest NO3 concentrations detected in the enhanced monitoring program. The oxidation of N species in HHF aquifers could also act as a mild source of acidity, due to the production of 2 moles of H+ for every mole of NH3.
The pH of all groundwaters continues to be mildly acidic at values around 4 to 5. The pH appears to be influenced by the presence of high concentrations of dissolved carbon dioxide (CO2 (aq)), especially at bore 3135U which consistently had the highest CO2 (aq) concentrations measured in the field. On the basis of the mixing analysis in Section 3.5.7, however, the continuing mild acidity of HHF groundwaters cannot be explained by this process alone. The oxidation of the NH3 released by organic degradation may be a minor source of acidity. Further mechanisms for acidity will be discussed later in this chapter.