Reclamation of acid waters using sewage sludge

Environmental Pollution 57 (1989) 251-274

Reclamation of Acid Waters using Sewage Sludge

W. Davison, C. S. Reynolds, E. Tipping Freshwater Biological Association, Windermere Laboratory, The Ferry House, Ambleside, Cumbria LA22 0LP, UK

& R. F. Needham Hepworth Minerals and Chemicals plc, Brookside Hall, Congleton Road, Arclid, Sandbach, Cheshire CW11 0SS, UK

(Received 10 June 1988; revised version received 16 September 1988; accepted 22 September 1988)

ABSTRACT An exhausted sand quarry which had filled with acid water (pH 3)from the oxidation of pyrite was treated with calcium hydroxide to neutralize the water (pH 8), and sewage sludge to prevent further ingress of acid. The water remained neutral for 2 years, an appreciable quantity of base being generated by the reduction of sulphate to sulphide in the anoxic sediment formed by the sewage sludge. After this time the water reverted to acid conditions, chiefly because the lake was too shallow to retain the sewage sludge over a sufficiently large area of its bed. Incubation experiments showed that the sewage sludge had a large capacity for sulphate reduction, whieh was equally efficient in acid or neutral waters and that the areal rate of consumption was sufficiently fast to neutralize all incoming acid, if at least 50% of the lake bed was covered with sludge. Throughout the course of the field investigations there was no foul smell and the lake was quickly colonized by phytoplankton, macrophytes and insects. Although nutrients associated with the sewage sludge stimulated photosynthesis and so caused the generation of additional organic matter, they were exhausted within two years. To ensure permanent reclamation, phosphate fertilizer could be added once the initial supply has been consumed. 251 Environ. Pollut. 0269-7491/89/$03.50 '9 1989 Elsevier Science Publishers Ltd, England. Printed in Great Britain

252

W. Davison, C. S. Reynolds, E. Tipping, R. F. Needham Neutral&ation removed trace metals from the system, presumably due to formation of insoluble oxyhydroxide and carbonates. The solubility of aluminium was apparently controlled by a basic aluminium sulphate (jurbanite).

INTRODUCTION Concern about the acidification of lakes has stimulated a series of studies which have shown that sediments may have an appr~iable capacity for generating base (Kelly et aL, 1982; Baker et al., 1985; Cook et al., 1986; Schindler et al., 1986; Davison, 1987). Alkalinity is produced when NO 3, MnO 2, FeOOH and SO ] - are used as terminal electron acceptors in the microbially mediated oxidation of organic matter. The principle of base generation from decomposing organic matter is not new and has been suggested as a potential pollution abatement procedure for acid mine water (Tuttle et al., 1969). Such waters contain appreciable concentrations of sulphuric acid which result from the oxidation of pyrite (Stumm & Morgan, 1981). Once the supply of acid ceases, the long-term recovery of strip mine lakes depends on the availability of organic material, and so King et al., (1974) have suggested that the addition of organic material, such as sawdust, straw, leaves, newspaper, manure or sewage sludge, will accelerate recovery. They favoured placing the material in discrete piles rather than spreading it evenly over the lake bottom. Pyrite is a common mineral and any activity which modifies the landscape may expose it to oxygen and so induce acidity (eqn (1)). Extraction of sand 4 FeS2 + 15 O2 + 10H20 ---~4 FeO(OH) + 8 S O l - + 16H +

(1)

and gravel for industrial and construction purposes is widespread and, where the excavation extends below the water table, lakes and ponds are formed. If the water is found to be acid, reclamation is complicated because free hydraulic exchange with the surrounding groundwater provides a continuing source of acidity. Treating such waters with lime would ameliorate the situation, but only temporarily because the flow of groundwater would eventually recreate acid conditions. We proposed a method for producing a self-regulating system which is capable of continually neutralizing the ground water, so maintaining a stable pH. In a pilot study to test the scheme, we used a combined treatment, adding lime to raise the pH, and organic matter to act as a redox buffer and provide continuing stability. Formation of a uniform layer of organic matter over the surface of the sediment of the lake combats acidity in several ways. Decomposition

Reclamation 0[' acid waters using sewage sludge

253

processes lead to reducing conditions, so that sulphate is converted to sulphide, which is removed as iron sulphide (eqn (2)). This reducing zone, situated at the 9 C H 2 0 + 4 FeO(OH) + 4 S O ] - + 8 H + ~ 4 FeS + 9 CO 2 + 15 H 2 0 (2j sediment-water interface, acts as a 'chemical filter', removing sulphate as it enters the lake via the groundwater. The large area of reducing sediment also helps to prevent oxygen from coming into contact with the groundwater and so slows down the production of sulphate from oxidation of pyrite. If the organic carbon is added as sewage sludge, nutrients are introduced which encourage primary productivity. Thus, additional carbon, originating from atmospheric CO2, is continually added to the lake and eventually accumulates as organic-rich sediment. Photosynthesis is effectively used to combat acidification and ultimately a new steady state pH will be established. We report here the detailed results of a long-term pilot study to test the scheme: some preliminary results have already been reported (Davison, 1986).

EXPERIMENTAL Site

Hepworth Minerals and Chemicals owns a 1500-acre estate at Leziate, near King's Lynn in Norfolk, England, from which high grade silica sands have been excavated for over a century. The excavations extend below the water table and so, when dewatering ceases, lakes are formed. As the pH of the water is below 3, there is a requirement for a permanent neutralization scheme. One of the smaller lakes, Blue Lagoon, was used in a pilot scheme to test the feasibility of using a combined treatment of lime and sewage sludge. Blue Lagoon had remained undisturbed for over 10 years and the water had a pH near to 3. The lake had an area of 3-6 ha, a volume of 52,000 m 3, a mean depth of 1.5 m and a maximum depth of 4"5 m. As it was literally a hole in the sand there were no inflows and a negligible catchment. Treatment

On 3 April 1984, 477,000 litres of sewage sludge containing 4"2% solids (78"6% organic volatile) were pumped into the lake via a pipeline which was moved in an arc to cover approximately one quarter of the lake area. Wind action was expected to distribute the sludge throughout the rest of the lake.

254

W. Davison, C. S. Reynolds, E. Tipping, R. F. Needham

On 30 April 1984, 16.98 tonnes of high calcium hydrated lime, equivalent to 15-3 tonnes of calcium hydroxide, were added via a submerged pipe near to the deepest point.

Sampling and analysis Water samples were collected at approximately two or four weekly intervals from the edge of the lake by plunging a container below the surface, removing its cap and allowing it to fill. The cap was returned before removing the container from the lake. Care was necessary to avoid disturbing the bottom of the lake and including some particulate material in the sample. Three polythene containers were used. One was covered with black plastic tape and was used for determining air-equilibrated pH, another was primed with Lugol's iodine solution and was used for subsequent algological analyses, and the third container was used for all other routine chemical analyses. Samples were analyzed within 2-4 weeks. Some additional samples, from a depth of 3 m, were collected using a Ruttner sampler from a boat at the deepest site. On these occasions, in situ oxygen was measured using a Partech Electronics probe. Extra pH measurements were made locally by collecting 250-ml water samples from different sites round the edge and measuring them in a laboratory within 20 min of collection. Routine analyses of the waters were performed using well-established procedures (Mackereth et al., 1978). Calcium, sodium, magnesium and potassium, were determined by atomic absorption spectroscopy on unfiltered samples. Nitrate, ammonia, phosphate, total phosphorus, silica, aluminium and total iron were determined colorimetrically. Turbidimetry or ion chromatography was used for sulphate, titration for alkalinity, titration or ion chromatography for chloride, and conventional electrochemical systems for pH and conductivity. Copper, lead, cadmium, zinc and manganese were measured using either differential pulse polarography or atomic absorption spectroscopy. Total phosphorus and conductivity were measured on unfiltered samples. Analyses for ammonia, sulphate, chloride, alkalinity, nitrate and silica were performed on unfiltered samples up to and including 26 September, 1984, when possible contamination from disturbed sediment during collection was recognized as a problem. Thereafter, measurements were made on filtered samples. All dissolved phosphate and total iron measurements were made on filtered samples (Whatman GF/C). Air-equilibrated pH was measured after bubbling air through the sample for at least 30 min. The resulting values were assumed to reflect the pH of the

Reclamation of acid waters using sewage sludge

255

water free from the influences of deficiencies or excesses of carbon dioxide, brought about by photosynthesis or respiration. For algological evaluation, a portion of the iodine 'fixed' water sample was transferred to a sedimentation tube where the algae were allowed to settle by gravity. The number of cells and approximate mean volumes of each algal species were determined.

Sediment samples On 27 August 1985, the depth of sludge at various sites in the lake was assessed by visually examining sediment cores collected with a Jenkin sampler (Ohnstad & Jones, 1982). Four cores of sediment, complete with overlying water, were collected from the deepest site and retained for subsequent analysis. Two of them, BLC and BLE, were sectioned, and acid volatile sulphide (Davison & Lishman, 1983), carbon, hydrogen and nitrogen (Carlo Erba 1106 elemental analyser), and dry weight were determined at 1-cm intervals. Another two, BLA and BLB, were fitted with sealed syringe sampling systems and incubated in the dark at 17°C. Five-millilitre samples of the overlying water were withdrawn at weekly or fortnightly intervals for sulphate analysis, the balance being made up with deoxygenated Blue Lagoon water. After depletion of the sulphate, most of the overlying water was withdrawn under anoxic conditions and replaced with fresh, deoxygenated Blue Lagoon water. Incubation of BLA and BLB was then continued at 10°C and 4°C, respectively. Further incubations were carried out over a 2-year period, after which the sediment was extruded and analyzed. On 25 February, 1987, further sediment cores were collected for analysis and incubation.

R E S U L T S A N D DISCUSSION

General chemistry The sewage sludge which was added settled very quickly so that within a few days the water of the lake looked clear. Examination of sediment cores on 27 August 1985 revealed that little sludge was retained on the sediment in shallow ( < 2 m) water, but that most of the deep site ( > 4 m of water) was well covered, with up to 7 cm of sludge. Adding sewage sludge had no effect on the pH, but within a few days of adding calcium hydroxide the in situ pH rose rapidly to values approaching

W. Davison, C. S. Reynolds, E. Tipping, R. F. Needham

256

10, and then quickly (days) declined to near 8 as carbon dioxide dissolved from the atmosphere. Calcium concentrations were temporarily very high, before declining as, according to the equilibrium calculations, calcium carbonate became the phase controlling solubility (Fig. 1). For the first two years, the air-equilibrated pH (Fig. 1) remained neutral, but progressively declined from 8 to 6.5. The pH then dropped rapidly to c. 5, and then more slowly, to approach the starting pH by September 1987. A

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Temporal variation of pH, calcium, alkalinity and conductivity.

Reclamation of acid waters using sewage sludge

257

rapid decline in pH from 6.5 to 5 is consistent with the titration of an airequilibrated carbonate solution with a strong acid, the buffer capacity being at a minimum in this region (Stumm & Morgan, 1981) where alkalinity is near to zero. For two years, the concentrations of K, Mg, Na and C1 varied seasonally (Fig. 2), with a maximum in late summer that may be associated with the effects ofevapo-transpiration. In early 1986, major changes were apparent in their concentrations, particularly noticeable in magnesium and sulphate. At this time, active dewatering commenced in an adjacent excavation within 50 m of Blue Lagoon causing a 0"5-m lowering of the water level. The rapid increase in the concentrations of sulphate and magnesium indicates that a new, more acid groundwater may be entering. The lake water was always uncoloured, samples collected on 8 January 1986 and 10 July 1986 having 2"7 and 0.6mg litre-~ of dissolved organic carbon.

Calcium and alkalinity The concentration of calcium increased markedly immediately after hydrated lime was added in the initial neutralization. Excess lime was added: analysis of the water revealed that 3-2 tonnes of calcium initially dissolved, while 5"1 tonnes remained in the solid phase. In theory the excess lime could continue to dissolve as more acid entered the lake. In practice there was no evidence for further dissolution, which is consistent with the work of Svedrup et al. (1984), who showed that the stagnant sediment boundary layer, and development of inorganic and organic surface coatings, inhibits dissolution of material on the bottom of a lake. Assuming that no further dissolution of added calcium occurs, it is possible to model the change in concentration with time. The lake is assumed to be a continuously stirred tank with an inflow of water containing calcium at concentration Cain, which is the value measured in the lake before lime was added. A plot of log ( C a c o r r - Ca~,) versus time should give a straight line with a slope of the reciprocal of 2.3 x the mean hydraulic residence time of the water. Cacorr is the measured calcium concentration in the lake, corrected for evaporation or dilution effects by normalizing with respect to the concentration of sodium which is assumed to be conservative. Although there is considerable scatter in the data, it approximates very well, for a period of almost four years, to a straight line (Fig. 3), with a slope which corresponds to a residence time of 3"8 years. The good fit to a theoretical model of a straight line has several implications. There is little evidence for any change in residence time during this period, so that nearby dewatering which commenced during early 1986

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must have modified the source and hence chemical composition of the groundwater, rather than increasing the flow through the lake. Progressive clogging of the pores of the sand by the fine organic material does not appear to have resulted in any pronounced decrease in the flow of water. Little of the solid phase calcium could have dissolved after the initial neutralization because, for a straight line to result, the dissolution would have had to occur at a constant rate, which would be unlikely. After neutralization, further dissolution of calcium hydroxide produces alkalinity, and so if no other mechanism operates, alkalinity could be expected to decline in a similar way to calcium. Clearly it does not (Fig. 1); in particular it progressively increased during most of 1985 and the increase could not be attributed to concentration by evapo-transpiration. By early 1986, the generally more rapid decrease in the concentration of calcium could be quantified so that the difference amounted to 1 m eq litre '. Thus, during this time 1 m eq litre- 1 of alkalinity must have been generated by other mechanisms within the lake. One such mechanism is the reduction of ferric iron and sulphate, by the decomposition of organic matter, to form ferrous sulphide. From eqn (2) we know that approximately 1 g of organic matter is capable of consuming 0.0296 moles of acid (Davison, 1987). The theoretical maximum possible production of alkalinity of 9"0m eq litre 1 was calculated from the amount of organic material added, which was equivalent to 0"30 g of dry organic material for each litre of wa~er. Theretbre anoxic reduction of 11% of the organic material over a two-year period with concomitant consumption of acid, could account for the observed alkalinity changes. This internal generation of 1 m eq litre- 1 of base is considerably

260

W. Davison, C. S. Reynolds, E. Tipping, R. F. Needham

more than that produced by the initial dissolution of calcium hydroxide (c. 0.6 m eq litre- x, Fig. 1), and so it must have been important in maintaining a neutral pH during the first two years after treatment.

Sulphate reduction For the sewage sludge to generate base, sulphate must be reduced in the sediment according to eqn (2). The initial idea was that this organic layer would intercept sulphate seeping into the lake from the groundwater, and by reduction, neutralize this acid before it became part of the main water body of the lake. Thus, neutralization of incoming acid does not require active reduction of sulphate already within the lake, but, if the incoming water contained less acid, the sulphate concentration in the lake would gradually decline according to the mean hydraulic residence time of the system. The sulphate concentration progressively decreased from its value immediately after neutralization until early 1986, when fresh supplies were introduced as a consequence of nearby working. A plot of the sulphate concentration, normalized with respect to the concentration of sodium to correct for evapotranspiration and dilution effects, gave a straight line (Davison, 1986) which declined systematically from 130 mg litre- 1 in early 1984 to 110 mg litre - 1 in early 1986. This consumption of sulphate of 20mg litre -1 could be expressed as 0.42 m eq litre-1 of sulphuric acid. A lower value than the estimate of the internally generated base, which was 1 m eq litre- ~ (see preceding section), is to be expected, as it is based on reduction of sulphate from the water column and substantially excludes reduction of incoming acid from groundwater. Analysis of sediment collected from the deepest part of the lake also showed that sulphate was being reduced. Cores collected from the deepest point in August 1985 (BLC, BLE) had high concentrations of acid volatile sulphur, AVS (Table 1), which provides a measure of amorphous iron sulphide, in the organic-rich layers of sediment overlying the sand. Cores collected from the same site in February 1987 (BLX, BLZ), when the lake waters had become acid again, had slightly lower values of organic carbon, probably as a result of recent re-landscaping of the adjoining banking, which had introduced a clearly observable layer of fresh sand onto the surface of the sediments. Nevertheless, except near to the sediment surface, high levels of AVS were still present. Reduction processes within organic-rich sediments buffer the pH to near-neutral values even when the overlying waters are acid (Herlihy & Mills, 1986). However, acid water at the sediment-water interface would dissolve FeS or prevent its formation. Two cores containing sediment and water, BLA and BLB, collected in August 1985, were sealed and incubated anoxically in the dark. Sub-samples

Reclamation o f acid waters using sewage sludge

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TABLE I AnalysisofSedimentCores Depth (crn) TotalS% BLC BLE BLX BLZ BLA

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0"018 <0.001 0.044 0"11 0'60 0"072 0.006 1 <0-001 1.20 0.71

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5"8 6"7 7"4 4.0 19-0

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were analyzed for sulphate, and after exhaustion of the sulphate in the overlying water, the water was exchanged, while excluding air, with fresh deoxygenated water from Blue Lagoon. This incubation procedure was continued for more than two years at various temperatures (Fig. 4). Results from the two cores were similar throughout the period. Sulphate was removed at a rate of 835 mg m - 2 day- ~ in the initial incubation at 17°C and at 278 mg m - 2 day- ~ at 10°C and 4°C. Two years later the same sediments, but with an overlying water at pH 3.7, consumed sulphate at 553 mg m - 2 day- ~ when incubated at 20°C. A core collected in January 1987, BLY, and incubated with the same water at 20°C, consumed sulphate at a rate of 980 mg m - 2 day- 1. It appears that the 3-year-old sediment, overlain by acid water, can reduce sulphate just as quickly as the 1.5-year-old material overlain by neutral water, showing that sewage sludge is effective at consuming sulphate over a wide range o f pH, as has been found for natural lake sediments (Herlihy & Mills, 1986). Even after this prolonged incubation, the rate o f sulphate consumption has declined by less than a

W. Davis•n, C. S. Reynolds, E. Tipping, R. F. Needham

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factor o f two, showing that the organic decomposition processes which drive the reduction of sulphate can continue at a high rate for a long time. Assuming that the incoming groundwater has the composition of the lake water prior to treatment, it is possible to calculate, using 3-8 years for the mean residence time, the rate o f sulphate reduction required to maintain neutrality, if the whole of the bottom of the lake was covered with sewage sludge, as 74 mg m - 2 day - 1. This value is less than one third of the measured rate of sulphate reduction at 10 or 4°C of 228 mg m - z d a y - 1. Therefore, base generated by the anoxic decomposition of sewage sludge is theoretically able to cope with the ingress o f acid into this system, providing that the sewage sludge covers most of the lake bed.

Reclamation ~[ acid waters using sewage sludge

263

The sediment from core BLA was analyzed after completion of the longterm incubation. Concentrations of total sulphur were much higher than in similar cores, BLC, BLE, prior to incubation (Table 1). The extension of high values to greater depths may be due to BLA having a thicker organic layer, but similar values for organic carbon and C/N ratios indicate that BLC, BLE and BLA are all comparable. Summation of the sulphur from 0 to 8 cm in BLA gives 27"5mg cm-Z; 12mg cm -z greater than in BLC. This generation of reduced sulphur of about 12 mg cm-2 compares reasonably with an estimate, from sulphate concentrations in the topping up water, of the sulphate consumed during the incubation of BLC and BLE, of 8 mg cm -2. If the sulphur in the sediment was accumulating as amorphous FeS it would be detected as acid volatile sulphur. Surprisingly, this is not the case (Table 1); indeed AVS is depleted in the surface sediments of BLA, possibly due to the overlying water being acid or to an ingress of oxygen during the two-month period between completion of the incubation and analysis. Had oxidation occurred, the sulphur could have been present as sulphate, otherwise it must be either pyrite or elemental sulphur. Direct formation of pyrite in freshwater sediments in situations such as this has been previously suggested (Davison et al., 1985). Metals

Not surprisingly the concentration of metals was high in the initially very acid waters (Table 2, Fig. 5). After neutralization, the concentration of iron, aluminium, zinc and manganese declined rapidly, as insoluble hydroxides TABLE 2 C o n c e n t r a t i o n s o f Dissolved Trace Metals

Date

23.2.84 2.4.84 3.4.84 12.4.84 25.4.84 10.5.84 23.5.84 23.5.84 18.1.86 10.7.86

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83 42 42 20 42 54 75 83 2.5 0.2

56 < 5 < 5 < 5 < 5 < 5 < 5 < 5 <0.1 <0.1

373 226 226 242 236 18 < 5 9 24 175

205 205 185 205 205 < 20 < 20 < 20

IV. Davison, C. S. Reynolds, E. Tipping, R. F. Needham

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and carbonates were formed and co-precipitation and adsorption became possible. A residue of aluminium remained for some time, presumably due to soluble anions that are present in alkaline (pH > 8) solutions. When the lake had been neutral for sometime, in early 1986, the concentrations of trace metals were very low (Table 2), but as acid conditions returned so did the soluble metal ions, as exemplified by iron and aluminium (Fig. 6). The removal and re-dissolution of aluminium over wide ranges of pH (2.5-10), sulphate concentration (0.5-5 mm) and silicic acid concentration (0-01-2 m i ) provided an opportunity to consider the possible control of AI concentrations by solid phases. To do this, the chemical speciation of A1 was calculated for each Blue Lagoon sample in which dissolved A1 was non-zero, and for ten water samples collected from similar excavations on the same

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site. The calculations took into account ionic strength, the complexation of A13 + by OH -, SO 2- and dissolved humic substances, and the complexation of Ca 2 + by SO 2-, but ignored (due to lack of data) Al-fluoride complexation and competing reactions involving dissolved Fe. Calculations were performed for an assumed (representative) temperature of 10°C, using equilibrium constants and enthalpies tabulated by Tipping et aL (1988a) and

266

W. Davison, C. S. Reynolds, E. Tipping, R. F. Needham

by Smith & Martell (1976). A constant (low) concentration of humic substances of 2 mg litre- ~was assumed, on the basis of three determinations of dissolved organic carbon in Blue Lagoon samples, and model III of Tipping et al. (1988b) was used to estimate A13+ complexation by the humics; this had a minor effect on the A1 speciations for these samples, less than 10% of the dissolved A1 being humic-complexed in all cases. Ion activity products (IAPs) were calculated for the stoichiometries AI(OH)3 (gibbsite, or more amorphous aluminium hydroxides), AlzSi2Os(OH)4 (kaolinite, halloysite), A1OHSO4 (jurbanite) and [AI(OH)3] ~_x [SiO2]x (reversible metastable aluminosilicate, x - - 1 " 2 4 - 0 . 1 3 5 p H ; Paces, 1978). Phases with these compositions are generally considered the most likely regulators of A1 concentrations in natural waters (Nordstrom, 1982; Bache, 1986). Plots of logloIAP, or (log~oIAP - l'59pH) in the case of the metastable aluminosilicate (Paces, 1978), vs. pH for the 30 samples considered are shown in Fig. 6; note that the IAPs are formulated in terms of the activity of H +, not O H - . Considering first the pH range 2.5-6, which covers 26 of the 30 samples, it is apparent from Fig. 6 that the most likely controlling phase, judged by the constancy or otherwise of the log10IAP values, is jurbanite, A1OHSO 4. The mean log~oIAP value for this stoichiometry is -3.90, with a standard deviation of 0-21; this is close to the values of -3"8 and -3"2 (both at 25°C) estimated for the solubility product ofjurbanite by van Breemen (1973) and Nordstrom (1982), respectively. Clearly, there is no question of either gibbsite or kaolinite/halloysite controlling A1 concentration, since the log~oIAP values for both stoichiometries vary markedly between pH 2"5 and 6. The possibility of equilibrium with a metastable aluminosilicate phase cannot be ruled out entirely, although the value of (log~oIAP- 1"59 pH) does appear to vary systematically with pH. The mean (log~oIAP - 1-59 pH) value is 6"65 (standard deviation 0"50), similar to that of - 5-89 estimated by Paces (1978). At pH values greater than 6 (4 data points only) the water samples are calculated to have been substantially undersaturated with respect to jurbanite, and substantially oversaturated with respect to both kaolinite and halloysite, suggesting that the formation of either form of crystalline aluminosilicate was kinetically unfavourable in the waters under consideration. The mean logxoIAP value for AI(OH)3 (10-5) is in agreement with effective solubility products estimated for freshly formed amorphous A1 precipitates (cf. Tipping et al. 1988c), and bearing in mind the large uncertainties in speciation calculations at higher pH values, the lOgl0IAP values are very approximately constant (Fig. 6). Again, however, the possibility of control by a metastable aluminosilicate cannot be ruled out. In summary, it appears that, on the basis of the present data, the best -

Reclamation o[acid waters using sewage sludge

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explanation of dissolved A1 concentrations in Blue Lagoon and neighbouring waters is a combination of control by jurbanite at pH < 6, and more tentatively by amorphous aluminium hydroxide at pH > 6.

Biological developments The effects of the treatments applied to Blue Lagoon upon its biotic communities were monitored qualitatively or at best semi-quantitatively. The most immediate effects were evident among the planktonic floral and faunal components. The development of rooted macrophytes and a more balanced trophic structure took relatively longer to establish. Semi-quantitative assessments of phytoplankton are based on dip samples collected from the edge of the lake and enumerated by the iodine sedimentation and inverted microscope technique of Lund et al. (1958). The data thus obtained are subject to several sources of error; these include those inherent in sub-sampling and counting, but particularly the possibility that the plankton communities are distributed heterogeneously in the water and so the sample may be atypical. Thus each point in Fig. 7, which shows the fluctuations in the concentrations of live algal biomass, carries, according to accumulated experience over many years, an error margin of about _+50%. The main features are still clear. Phytoplankton populations increased significantly following treatment in 1984, but, after a sharp decline in 1985, they remained at a generally low level. Whereas in 1984 phytoplankton biovolume was frequently within the range of 8-12 mm 3 litre- 1 associated with eutrophic lakes, it was much less in 1985 (generally 0"5-1"0 mm 3 litre ~i, 1986 ( < 2 - 0 m m 3 litre -1) and 1987 ( < 0 " 1 3 m m 3 litre-lj. The type of algae present and their relative abundance and dominance appeared to be related to the different stages of the treatment. Prior to any treatment there were only a few acid-tolerant species of Chlamydomonas and Chlorella. Adding sewage sludge stimulated a rapid increase in these algae (especially of the Chlamydomonas), and of large numbers of presumably acid-tolerant Sphaerella sp. and of an Ochromonad (Chrysophyceae). The addition of lime, a month later, was followed by a collapse in the existing phytoplankton and the development of populations of diatoms (Synedra nana), a different species of Chhtmydomonas and, later, by Scenedesmus and Chlorella spp. These latter are typical of alkaline, sewage lagoons and attest to the remarkable modification of the water of Blue Lagoon. The much reduced biomass during 1985 did not immediately represent a reversion to the earlier acidophilic plankton, nor to one necessarily limited by lack of nutrients, for the dominant species {Scenedesmus spp, Qvclotella pseudostelligera and Rhodomonas minuta) remained those of productive, nutrient-rich small lakes. It was more likely a further consequence of the lake

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Reclamation o] acid waters using sewage sludge

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recovery and the prolific growth of macrophytes (see later). In 1986, the samples continued to be dominated by small-celled nanoplanktonic forms of Chlorella, Chlamydomonas and Rhodomonas, although small diatoms (Navicula, Synedra and Cyclotella spp.) were relatively numerous in April and May. A substantial population of Dinobryon was abundant in May-June. Many other species observed in 1984 and 1985 were recorded briefly or sporadically in 1986. Not until 1987 did the species composition and standing crops revert substantially to those observed prior to treatment. Measurements of coliform bacteria were made on four samples collected on 19 November 1985. They each yielded values of < ten Escherichia coli cells per 100ml showing that the water was free from this normal faecal bacterial indicator. Planktonic animals were noted in the water samples for phytoplankton, providing qualitative information about the colonization of the lake. Prior to treatment there was no evidence of animals. Small rhizopods and then ciliates (cf. Colpidium) became established during June and July, 1984; several species of planktonic rotifer (including Ploesoma, Polyarthra, Syncheata) also developed at the time of the Scenedesmus maximum. Daphnia spp. were recorded in a single (September) sample. During 1985, Coleps (Ciliophora) and Keratella spp. (Rotifera) were in evidence, while, during June and July, a substantial population of the planktonic calanoid. Diaptomus gracilis, flourished in the Lagoon. Its demise coincided with the introduction of 500 rainbow trout (Salmo gairdneri) on 1 July 1985. When these trout were sampled on 7 October 1985, their average length had increased from 12"5 to 30"5 cm and their average weight from 160 g to 330 g. This excellent growth, comparable to fish farm targets, was testament to the abundant food supply in Blue Lagoon. In the subsequent years, only occasional ciliates and rotifers continued to be encountered. This may have been attributable to inadequate sampling, though, in general, it would appear that the algal biomass became too sparse to support prolific grazing populations. Most of the rainbow trout were eventually removed by anglers. Once the pH had declined to 4"5, they would not have been expected to survive, although a single large specimen, in excess of 30 cm in length, was clearly seen on 26 September 1986 when the pH was 4.2. The macrophyte flora underwent a most spectacular change during the eighteen months following treatment. In early 1984, there was very little plant life in or around the water, apart from some rather stunted mosses, identified as Fontinalis. During the summer, annual meadow grass (Poa annua) began to seed in the strand lines around the shores and a number of plants ofPotamogeton natans became established in the water. In 1985, there was a luxuriant growth of Zannichellia palutris (with lesser amounts of Potamogeton natans and Ranunculus cf. aquatalis) which, by late August, had

270

W. Davison, C. S. Reynolds, E. Tipping, R. F. Needham

colonized most of the area covered by sewage-based organic sediment and had generated a dense underwater forest of macrophytes. Macrophyte development was extremely poor in 1986 and 1987. On 10 July 1986, some shoots of Potarnogeton natans were still growing, apparently from old roots, but there was no sign of Zannichellia palutris. Concentrations of dissolved phosphate rose dramatically from their very

30

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Reclamation o f acM waters using sewage sludge

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low background levels when sewage sludge was added (Fig. 8). The subsequent rapid decline, coincident with increased levels of total phosphorus, is consistent with the rapid growth of phytoplankton. Some of the very high levels of total phosphorus at this time were known to be due to entrainment of particles of bottom sediment in the samples. After this initial input, dissolved phosphate was rarely above the detection limit (0-6 or 3/~g litre 1, depending on the method). Phosphorus evidently limits algal growth, and there appears to be no long-term release of phosphorus from sewage sludge. The rapid decline in phytoplankton populations (Fig. 7) and macrophytes after the first two years is most probably due to limitation by phosphorus. Although the ammonia and nitrate nitrogen results (Fig. 8) are subject to error due to the possibility ofinterconversion during storage, they show that nitrogen was never limiting. Nitrate was severely depleted in the late summer of 1985, but there were still reserves of ammonia, and in subsequent years there was little evidence for depression of either nitrogen species. One of the envisaged benefits of the scheme was that nutrients from the sewage sludge would stimulate photosynthesis and add to the organic carbon on the lake bed. Whereas this was true initially, there is an insufficient long-term supply of phosphorus. It is possible to calculate how much phosphate would be required annually to generate sufficient carbon from primary productivity to neutralize the incoming acid, assuming all the carbon anoxically decomposed to produce iron sulphide (Davison, 1987). Twenty-six kilograms of phosphorus would be required, equivalent to 117 kg of Na 2 HPO4 which could be added as fertilizer. Whether this amount of phosphate would have prevented a return to acidity is uncertain. The assumption that all the organic carbon would be degraded anoxically is optimistic. Some must undergo oxic decomposition, and indeed during June, July and August in 1984 the concentrations of dissolved oxygen in the deep (4 m) water were depleted. Conversely, some of the phosphate will be recycled back into the lake as carbon is decomposed. The effectiveness of adding phosphate as fertilizer to combat acidity (Yan& Lafrance, 1984) needs to be further tested in field situations.

G E N E R A L DISCUSSION A concern at the start of this work was the possible harmful effects of sewage sludge, such as the production of noxious smells or eutrophication. Neither effect occurred, probably because this work involved adding a single dose of sewage sludge. Foul smelling or eutrophic systems are usually continuously supplied with organic material or nutrients.

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W. Davison, C. S. Reynolds, E. Tipping, R. F. Needham

Consideration of chemical changes showed that base was being generated internally in the system via a process of sulphate reduction. Incubation experiments have shown that the areal rate of sulphate and acid consumption is, in principle, able to neutralize the additional acid which is continually supplied from the groundwater. Failure to maintain a neutral pH beyond two years can be attributed to several factors, but chiefly it is due to the lake being too shallow, allowing wind and wave action to scour the sewage sludge from all but the deepest parts of the bed. The success of the scheme depends on a stable covering of sewage sludge over most (at least half) of the bottom of the lake. This can be realized only for lakes with an appropriate morphometry. Ideally, they should be well sheltered and flat bottomed with steeply sloping sides and a mean depth in excess of 5 m. In phosphate-limited systems, such as Blue Lagoon, adding only sewage sludge will not provide a permanent buffer. The organic matter that is added provides a finite reservoir of acid neutralizing capacity which is enhanced by material supplied from photosynthesis within the lake. Nutrients released from the sewage sludge stimulate photosynthesis, but for Blue Lagoon this supply was exhausted within two years. Such a deficiency could be remedied by adding phosphate as a soluble salt. On a weight per weight basis phosphate is potentially 33 times more efficient at generating base than calcium carbonate (Davison, 1987) Moreover, phosphate has the potential to be recycled within the lake as it is released from decomposing organic matter. Such re-cycling will never be 100% efficient, but it will considerably lessen the quantities of fertilizer which have to be added to maintain neutrality. It remains to be seen whether nitrate would also be required, or whether the lake would be colonized by cyanobacteria, capable of fixing nitrogen from the atmosphere (Pick & Lean, 1987). During the transition in conditions, from acid to neutral and then returning to acid, the chemistry of the lake behaved predictably. In particular, the concentrations of trace metals, including iron and aluminium, were high in acid waters and low in neutral conditions, due to solubility considerations. The relationship between pH and the concentrations of aluminium and sulphate were used to show that jurbanite was the phase controlling the solubility of aluminium.

ACKNOWLEDGEMENTS Hepworth Minerals and Chemicals and the Natural Environment Research Council provided financial support. We thank the staff of H M C and the Freshwater Biological Association for their enthusiastic help with the project.

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REFERENCES Bache, B. W. (1986). Aluminium mobilization in soils and waters. J. Geol. Soc. 143, 699-706. Baker, L. A., Brezonik, P. L., Edgerton, E. S. & Ogburn, R. W. (1985). Sediment acid neutralization in softwater lakes. War., Air, Soil Poll., 25, 215-30. van Breemen, N. (1973). Dissolved aluminium in acid sulphate soils and in acid mine waters. Soil Sci. Soc. Am. Proc., 37, 694~7. Cook, R. B., Kelly, C. A., Schindler, D. W. & Turner, M. A. (1986). Mechanisms of hydrogen ion neutralization in an experimentally acidified lake. Limnol. Oceanogr., 31, 134-48. Davison, W. (1986). Sewage sludge as an acidity filter for groundwater-fed lakes. Nature, 322, 820-2. Davison, W. (1987). Internal elemental cycles affecting the long-term alkalinity status of lakes: Implications for lake restoration. Schweiz. Z. Hydrol., 49, 186-201. Davison, W. & Lishman, J. P. (1983). Rapid colorimetric procedure for the determination of acid volatile sulphide in sediments. Analyst, 108, 1235, 1239. Davison, W., Lishman, J. P. & Hilton, J. (1985). Formation of pyrite in freshwater sediments: Implications for C/S ratios. Geochim. Cosmochim. Acta 49, 1615-20. Herlihy, A. T. & Mills, A. L. (1986). The pH regime of sediments underlying acidified waters. Biogeochemistry, 3, 377-81. King, D. L., Simmler, J. J., Decker, C, S. & Ogg, C. W. (1974). Acid strip mine lake recovery. J. War. Poll. Contr. Fed., 46, 2301-15. Lund, J. W. G., Kipling, C. & Le Cren, E. D. (1958). The inverted microscope method of estimating algal numbers and the statistical basis of estimation by counting. Hydrobiologia, 11, 143-70. Mackereth, F. J., Heron, J. & Tailing, J. F. (1978). Water Analysis, No. 36. Scientific Publications of the Freshwater Biological Association. Nordstrom, D. K. (1982). The effect of sulphate on aluminium concentrations in natural waters: Some stability relations in the system A1203-SO3-H 20 at 298K, Geochim. Cosmochim. Acta, 46, 681-92. Ohnstad, F. R. & Jones, J. G. (1982). The Jenkin Surface-Mud Sampler: User Manual. Occasional Publications of the Freshwater Biological Association No. 15. Paces, T. (1978). Reversible control of aqueous aluminium and silica during the irreversible evolution of natural waters, Geochim. Cosmochim. Acta, 42, 1487-93. Pick, F. R. & Lean, D. R. S. (1987). The role of macronutrients (C, N, P) in controlling cyanobacterial dominance in temperate lakes. New ZealandJ. Mar. Freshwat. Res., 21, 425-34. Schindler, D. W., Turner, M. A., Stainton, M. P. & Linsey, G. A. (1986). Natural sources of acid neutralizing capacity in low alkalinity lakes of the Precambrian Shield. Science, 232, 844-7. Smith, R. M. & Martell, A. E. (1976). Critical Stability Constants, Vol. 4, Inorganic Complexes, Plenum Press, New York. Stumm, W. & Morgan, J. J. (1981). Aquatic Chemistry (2nd edn), Wiley, New York.

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Svedrup, H., Rassmussen, R. & Bjerle, I. (1984). A simple model for the re-acidification of limed lakes, taking the simultaneous deactivation and dissolution of calcite in the sediments into account. Chem. Scripta, 24, 53-66. Tipping, E., Woof, C., Backes, C. A. & Ohnstad, M. (1988a). Aluminium speciation in acidic natural waters: Testing of a model for Al-humic complexation. Water Res., 22, 321-6. Tipping, E., Backes, C. A. & Hurley, M. A. (1988b). The complexation of protons, aluminium and calcium by aquatic humic substances: A model incorporating binding-site heterogeneity and macroionic effects, Water Res., 22, 597-611. Tipping, E., Woof, C., Walters, P. B. & Ohnstad, M. (1988c). Conditions required for the precipitation ofaluminium in acidic natural waters. Water Res., 22, 585-92. Turtle, J. H., Dugan, P. R. & Randles, C. I. (1969). Microbial sulphate reduction and its potential utility as an acid mine water pollution abatement procedure. Appl. Microbiol., 17, 297-302. Yan, N. D. & Lafrance, C. (1984). Responses of acidic and neutralized lakes near Sudbury, Ontario to nutrient enrichment. In Environmental Impacts of Smelters, ed. J. Nriagu, Wiley, pp. 457-521.