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Physicochemical properties and stability of activated sludge flocs under temperature upshifts from 30 to 45 °C
Journal of Colloid and Interface Science 281 (2005) 136–145 www.elsevier.com/locate/jcis
Physicochemical properties and stability of activated sludge flocs under temperature upshifts from 30 to 45 ◦ C F. Morgan-Sagastume, D.G. Allen ∗ Department of Chemical Engineering and Applied Chemistry, Pulp & Paper Centre, University of Toronto, 200 College Street, Toronto, Ontario M5S 3E5, Canada Received 21 March 2004; accepted 6 August 2004 Available online 21 September 2004
Abstract The impacts of temperature shifts from 30 to 45 ◦ C on the structural stability and surface charge of activated sludge flocs were assessed in four sequencing batch reactors (SBRs) treating pulp and paper mill effluent. The improvement in floc stability was tested by sludge magnesium enrichment in one SBR and by operating another reactor at a high sludge retention time (SRT) of 33 days. Floc stability was characterized by dissociation constants with solutions of CaCl2 , KCl, urea, and ethylenediamine tetraacetate (EDTA). Surface charge was assessed by cationic–anionic titration and metals concentrations were also determined. The temperature shift consistently caused an increase in the negative sludge surface charge from approximately −0.180 to −0.300 meq/g MLSS. Magnesium enrichment and a high SRT of 33 days promoted less negatively charged sludge, dampened the increase in negative sludge surface charge, and yielded structurally stronger flocs; however, sludge deflocculation still occurred. Manganese and iron appeared to be released by sludge under the temperature shift. It was concluded that the temperature shift deteriorates the flocculating physicochemical properties of the sludge and that better floc stability achieved by magnesium enrichment and a high SRT is not enough to stop deflocculation. Further research is required to clarify the origin of the increase in negative sludge surface charge, the role of metals, and the governing factors in sludge deflocculation under such temperature shifts. 2004 Elsevier Inc. All rights reserved. Keywords: Activated sludge; Floc stability; Surface charge; Temperature upshifts; Metals; Deflocculation; Dissociation constants
1. Introduction The capacity of sludge biomass to remain flocculated, to settle fast, and to compact determines the efficiency of solid– liquid separation in the activated sludge process. Although good sludge settling and compaction result in the separation of sludge biomass from treated effluent, poor sludge flocculation can cause increased levels of effluent suspended solids (ESS). Poor sludge flocculation is associated with weak floc structure of part of the biomass, with pinpoint flocs and with free-swimming cells in suspension. All these conditions lead to solids discharges, to the loss of biocatalyst used in pollutant degradation, and to the discharge of bioactive compounds adsorbed onto biomass. * Corresponding author. Fax: +1-416-971-2106.
E-mail address: [email protected] (D.G. Allen). 0021-9797/$ – see front matter 2004 Elsevier Inc. All rights reserved. doi:10.1016/j.jcis.2004.08.026
Weakening of bioaggregate structure and deflocculation appear to be common responses of suspended sludge biomass to transient operating conditions and environmental stresses, such as shifts in substrate loadings [1], in ionic strength [2,3], and in the concentration of dissolved oxygen [4–6] and toxic compounds [7,8] or metals [9]. Activated sludge deflocculation has also been shown to occur due to mesophilic–thermophilic temperature transients from 30–35 to 45 ◦ C [10]. Temperature shifts from 30–35 to 45 ◦ C in treatment plants of pulp and paper mills result from seasonal variations and the operation of batch units and shutdowns/start-ups in the upstream mill. Sludge deflocculation under such temperature shifts impairs treatment performance by increasing ESS levels and decreasing the removal efficiency of chemical oxygen demand (COD), as reported for sequencing batch reactors (SBRs) treating bleached kraft pulp and paper mill effluent [11].
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A means of preventing sludge deflocculation from happening is increasing the structural stability of sludge flocs. The factors determining sludge bioflocculation, and thus floc structure and stability, have been studied in the past few decades, as recently reviewed by Liss [12]. Sludge floc structure and stability are influenced by physicochemical and microbial factors. Physicochemical factors include divalent cationic bridging of EPS [13]; Fe(III) polymer bridging or flocculation via Fe(OH)3 [14,15]; polymer bridging, originally proposed by Pavoni et al., 1972 (see [13,16,17]); electrostatic double-layer repulsive and van der Waals attractive interactions [3]; alginate gel formation with Ca2+ [18]; hydrophobic interactions [19]; and steric interactions [20]. Microbial factors include relative amount, of filamentous organisms [21] and sludge age [22]. Due to the varying nature of sludge, one or several of these factors may play a predominant role in floc stability in a given type of sludge; however, quantifying these relative forces or interactions in sludge flocs is challenging and not yet feasible [23]. Improving sludge bioflocculation characteristics through the addition of cations can increase sludge settleability and also floc structural stability. Improved sludge settling characteristics (assessed by low sludge volume indices (SVIs)) and sludge dewatering (assessed by low capillary suction time (CST) and specific resistance to filtration (SRF)) can be achieved by adding divalent cations (Ca2+ and Mg2+ ) to the sludge. These divalent cations are considered to improve bioflocculation by cationic bridging [24]. Particularly important in improving sludge settling and dewatering are 2+ the ratio of monovalent (Na+ , K+ , NH+ 4 ) to divalent (Ca 2+ 2+ 2+ and Mg ) cations and the ratio of Ca to Mg in the feed. Ratios of monovalent to divalent cations greater than 2 and Ca2+ /Mg2+ ratios significantly different from 1 have been related to deteriorated floc properties [25,26]. Increased sludge Mg concentrations have been correlated with better sludge settleability and compressibility [27]. The replacement of Ca2+ in sludge by ion exchange with counterions such as H+ , Na+ , and K+ was observed to increase the specific resistance to filtration [18]. Sludge bioflocculation is governed by sludge surface properties, such as hydrophobicity, surface charge, and the composition of extractable extracellular polymeric substances (EPS) and of total polymers in the sludge [28]. A higher ratio of proteins to carbohydrates in EPS rather than the total amount of EPS and a less negatively charged and more hydrophobic sludge have been correlated with lower ESS and therefore to better sludge flocculation [22,29]. Also, increased cell hydrophobicity has been correlated with floc and granule formation; cell hydrophobicity is known to lead to bacterial adhesion [30]. A way of inducing a less negatively charged and more hydrophobic sludge, and therefore structurally stronger sludge flocs, would be by operating at higher SRTs (e.g., greater than 12 days), as demonstrated by Liao et al. [22,29]. Insights into the mechanisms controlling the structural stability of activated sludge flocs can be obtained by investi-
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gating the response of sludge flocs subjected to different environmental conditions. For example, Liao et al. [31] tested the effects of electrostatic interactions, ionic interactions, and the absence of hydrogen bonds on sludge floc stability. Activated sludge flocs were exposed to suspending solutions of varying ionic strength and cationic valence and with varying concentrations of ethylenediamine tetraacetate (EDTA) and urea. The goal of this work was to assess the changes in surface properties and structural stability of manipulated and nonmanipulated activated sludge flocs under temperature shifts from 30 to 45 ◦ C in sequencing batch reactors (SBRs) treating bleached kraft pulp mill effluent. Sludge cationic enrichment and high sludge ages of 33 days were assessed as ways to increase floc structural stability in activated sludge. Floc structural stability was evaluated by subjecting the sludge flocs to different suspending solutions with various ionic strength, cationic valence, and urea and EDTA concentration.
2. Experimental procedures 2.1. Sequencing batch reactors (SBRs) Four parallel sequencing batch reactors (SBRs) were operated to treat primary clarified bleached kraft pulp mill effluent for 80 days. Two operating strategies for improving floc structural stability were assessed: sludge cationic enrichment with Mg2+ and a high sludge retention time (SRT) of 33 days. Each strategy was tested independently in two SBRs (Mg-SBR and SRT-SBR). The four SBRs were operated in three 8-h cycles per day, with a reaction volume of 1.55 l and a hydraulic retention time of 14 h. Each 8-h cycle consisted of 25 min anoxic filling with mixing, 385 min reaction with mixing and aeration, 60 min settling, and 10 min discharge. The dissolved oxygen levels were above 2–3 mg/l and the pH was between 7.4 and 8.4 in the four SBRs. The raw mill effluent was transported at 4 ◦ C to our laboratory immediately after collection and then frozen at −20 ◦C. The wastewater was thawed as required (about 84 l/week) and conditioned with NH4 Cl (Mallinckrodt Inc., Paris, KY) and (NH4 )2 HPO4 (BDH Inc., Toronto, ON) to achieve a solubleCOD:N:P ratio of 200:5:1. The soluble chemical oxygen demand (SCOD) concentration in the feed was relatively constant in the four SBRs (reference SBR: 282 ± 38 mg/l; Mg-SBR: 284 ± 35 mg/l; SRT-SBR: 296 ± 39 mg/l; control SBR: 283 ± 52 mg/l) although it decreased by less than 100 mg/l over the time of operation due to storage. The biomass inoculum (0.5 l/reactor) was return activated sludge from the same treatment plant from where the wastewater was collected. Details of the experimental setup are reported elsewhere [10,11]. Sludge cationic enrichment was achieved by adding MgCl2 ·6H2 O (Caledon Laboratories Ltd., Georgetown, ON) to the prepared feed of one SBR (Mg-SBR) in order to
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Table 1 Similarities and differences in operating conditions among the four parallel SBRs Characteristic
Reference SBR
Mg-SBR
SRT-SBR
Control-SBR
Sludge retention time (SRT) (d) Temperature shift from 30 to 45 ◦ C Sludge manipulation
∼20 YES None
∼20 YES Mg enrichment
∼33 YES Higher SRT
∼20 NO, T = 30 ◦ C None
achieve an equivalent Ca2+ /Mg2+ ratio close to 1 (0.9 ± 0.1 as Ca/Mg) and an equivalent ratio of monovalent (Na+ ) to divalent (Ca2+ and Mg2+ ) cations close to 2 [2.1 ± 0.5 as Na/(Ca and Mg)]. The total Mg concentration in the prepared feed was increased from 0.9 to 3.7 ± 0.5 meq/l. The Ca concentration in the raw mill effluent was 3.4 ± 0.4 meq/l. Ninety-eight percent of the total of K and Na concentrations in the mill effluent was Na (288 ± 39 mg Na/l). One of the four SBRs was operated at an SRT of 33 days (SRT − SBR = 33 ± 9 days), and the other three SBRs were operated at an SRT of ∼20 days (control SBR: 20 ± 6 mg/l; Mg-SBR: 20 ± 7 days; reference SBR: 22 ± 13 days). The SRTs were maintained by wasting mixed liquor as a function of ESS and mixed liquor total suspended solids (MLSS) concentrations every 2 days. The sludge in two of the three SBRs operating at an SRT of 20 days was not manipulated, and one of these two reactors (reference-SBR) was subjected to a temperature upshift while the other one (control-SBR) was operated at a constant temperature of 30 ◦ C as a control. Table 1 summarizes the similarities and differences in operating conditions among the four SBRs. 2.2. Temperature upshifts The temperature shifts from 30 to 45 ◦ C in three of the SBRs (reference-SBR, Mg-SBR, and SRT-SBR) were conducted in sequence, as shown in Fig. 1., within 1.5 to 2.9 h of the reaction phase of an operating cycle. The sludge from the reference and control SBRs was mixed and redistributed between these two SBRs on day 30, before the first shift. The reactors’ temperature was monitored with 76-mm mercury thermometers. Deep-chamber water baths (reference, Mg-, and SRT-SBRs: 33 L, 1295 PC, VWR Scientific, Mississauga, ON; MeOH SBR: Precision Model 188, Precision Scientific Inc.) connected to water jackets surrounding each reactor were used to control the reactors’ temperatures. The accuracy of the temperature readings was ±1 ◦C. 2.3. Sludge surface charge, pH, and analysis of metals Sludge surface charge was determined by cationic– anionic titration [22,32], and reported as meq of anionic polymer/g MLSS. A 0.002-N hexadimethrine bromide (Polybrene) solution and a 0.001-N sodium-salt–polyanetholesulfonic-acid solution were used as the cationic and the anionic standards, respectively. The detailed protocol is outlined elsewhere [10]. Although a recent study [33] indicates that colloid titration tends to overestimate sludge sur-
Fig. 1. Temperature profiles in the four parallel SBRs showing sequential temperature shifts from 30 to 45 ◦ C.
face charge, this simple technique is still useful in comparing relative values of sludge surface charge, as in this study, before and after the temperature shift and among SBRs. The sludge surface charge measured in this was may reflect the total floc sludge charge rather than the sludge surface charge alone [33]. Samples of the influent, treated effluent, and mixed liquor from the reactors were sent to Maxxam Analytics Inc. (Mississauga, Ontario) for acid-extractable metals analysis. The concentration of metals in the solids corresponds to the difference between the mixed liquor and treated effluent concentrations, divided by the MLSS concentration. The metals determined were Al, Ba, Be, Bi, Ca, Cd, Cr, Co, Cu, Fe, Pb, Mg, Mn, Mo, Ni, K, Ag, Na, Sr, Ti, W, V, and Zn. The samples were acidified with HNO3 at collection, transported at 4 ◦ C, and analyzed by inductively coupled plasma mass spectroscopy (ICP-MS) based on the United States Environmental Protection Agency’s (EPA) protocol 6020. The detection limits of the analyses were 0.001 mg/l for Mn, Ba, and Be; 0.003 mg/l for Cu, Mg, Sr, Ti, V, Zn, and Cd; 0.005 mg/l for Cr, Co, and Fe; 0.006 mg/l for Mo; 0.007 mg/l for Ag; 0.01 mg/l for Ni; 0.025 for Al and Pb; 0.04 mg/l for Bi and Ca; 0.05 mg/l for W; 0.06 mg/l for Na; and 1 mg/l for K. MLSS, MLVSS, and ESS were measured based on the Standard Methods [34]. An approximate 80-ml sample of mixed liquor was taken from each reactor during the reaction phase for measuring pH using a digital pH electrode (DigiSense, Cole–Parmer Instrument Company, Niles Illinois), conductivity (Radiometer A/S, Copenhagen, Denmark) and for MLSS samples.
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2.4. Sludge floc stability Sludge floc stability was characterized by dissociation constants determined from plots of cumulative absorbance versus number of washings with different chemical solutions, as described by Zita and Hermansson [3] and Liao et al. [31]. The dissociation constants are defined as unit of absorbance/g MLSS/washing and quantify the detachment of solids from sludge flocs due to successive washings with solutions of calcium chloride (CaCl2 ; Fisher), potassium chloride (KCl; Sigma), urea (Sigma U-5128, St. Louis, Missouri), and ethylenediamine tetraacetate (EDTA, tetrasodium salt dihydrate, Fisher). Higher dissociation constants imply a weaker sludge floc structure. The dissociation constants were determined for different solution concentrations: 0.5, 0.005, and 0.0005 M for CaCl2 ; 0.3, 0.005, and 0.0005 M for KCl; 2 and 8 M for urea; and 20 and 300 mg/l for EDTA. The absorbance was measured at 450 nm, and four or five washings were conducted with each sample. 2.5. Statistical analyses The statistical significance of differences between means from the same SBR before and after a temperature shift was assessed by using paired t-tests. The statistical significance of differences between means from two different SBRs and between means for unequal numbers of observations before and after the shift from the same SBR was assessed by ttests for independent samples and unequal variances. The statistical significance of differences between single observations and means was assessed by t-tests about the population mean (µ, assumed to be equal to the single observation) when the population standard deviation (σ ) was unknown. The statistical significance was assessed at the 95% confidence level, unless indicated, and the levels of significance (p) are reported for those cases where a significant difference was found.
3. Results 3.1. Sludge surface charge The temperature shifts from 30 to 45 ◦ C caused the sludge surface to become more negatively charged (Fig. 2). Statistically significant increases in the negative sludge surface charge were measured a few days after the shifts in the MgSBR (148% or 2.5-fold) from −0.068 ± 0.040 (day 38) to −0.169 ± 0.049 meq/g MLSS (day 45), in the reference SBR (91%) from −0.176 ± 0.059 (day 32) to −0.337 ± 0.012 meq/g MLSS (day 45), and in the SRT-SBR (22%) from −0.166 ± 0.005 (day 50) to −0.202 ± 0.001 meq/g MLSS (day 58). The more negative sludge surface charge resulting from the shift from 30 to 45 ◦ C is reproducible and consistent with our previous long-term experiments in which the sludge surface charge changed from −0.15 to
Fig. 2. Sludge surface charge before and after the temperature shift from 30 to 45 ◦ C in the four parallel SBRs. The error bars correspond to standard deviations from two samples. Arrows and labels in the figure indicate times when temperatures were shifted for various SBRs.
−0.24 meq/g MLSS due to a temperature shift from 35 to 45 ◦ C [11]. The sludge surface charge in the Mg-SBR and in the SRT-SBR before and after the shift remained less negative (>−0.202 meq/g MLSS) than that of the sludge in the reference SBR after the shift (<−0.240 meq/g MLSS). The sludge surface charge in the control SBR at a constant temperature of 30 ◦ C remained relatively constant and with values similar to those of the SRT-SBR. Mg enrichment, SRT∼33 days, and constant-temperature operation (control SBR = 30 ◦ C) promoted a less negatively charged sludge in comparison to the reference SBR (Fig. 2). The mixed liquor pH remained relatively constant in the four SBRs during the 80-day period of operation (reference SBR: pH 8.0 ± 0.2; Mg-SBR: pH 8.0 ± 0.2; SRT-SBR: pH 7.8 ± 0.2; control SBR: pH 7.9 ± 0.2), and was not significantly affected by the temperature shifts. No difference in the conductivity of sludge supernatants was detected before and after the shifts in the four SBRs. The addition of MgCl2 kept the mixed liquor conductivity in the Mg-SBR higher, as expected (reference SBR 1695 ± 120 µS/cm; Mg-SBR 2005 ± 199 µS/cm; SRT-SBR 1695 ± 108 µS/cm; control SBR 1701 ± 109 µS/cm). 3.2. Sludge metal concentrations The most abundant metal from those determined in the sludge was Ca (∼17–27 mg/g MLSS), followed by Na (∼6–13 mg/g MLSS) or Al (∼7–9 mg/g MLSS), Mg (∼3–4 mg/g MLSS), Mn (∼2–3 mg/g MLSS), Fe (∼2–4 mg/g MLSS), and K (∼2 mg/g MLSS). The sludge Mg concentration was approximately two- to threefold higher in the Mg-SBR (8.95 ± 2.32 mg/g MLSS) than in the other SBRs (reference SBR: 3.99 ± 1.64 mg/g MLSS; SRT-SBR: 3.53 ± 0.38 mg/g MLSS; and control SBR: 3.16 ± 0.31 mg/g MLSS). The original Mg concentration in the sludge solids of the Mg-SBR increased from 5.7 mg/g MLSS in the inoculum sludge to 7.2 mg/g MLSS in the
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sludge solids of the mixed liquor in the first 40 days of operation, due to the addition of Mg to the reactor’s feed. Zn and Ba concentrations in the sludge ranged from 0.1 to 1 mg/g MLSS, and the concentrations of Sr, Ti, Cu, Pb, Cr, and V were below 0.1 mg/g MLSS in the four SBRs. Ag, Be, Bi, Cd, Co, Cr, Mo, Ni, and W were not detected in the mixed liquor samples. The average concentrations of Na (∼272–322 mg/l), Ca (∼68–74 mg/l), Mg (∼10 mg/l), K (∼7–9 mg/l), and Al (∼2 mg/l) in the feed to the reactors were the highest from the metals analyzed, particularly Na, with concentrations of approximately 300 mg/l. The concentrations of Zn, Ba, Ti, V, and Cu in the feed were below 0.1 mg/l. The feed metal concentrations were relatively constant during the period of operation and among reactors. Most of the concentration of metals in the influent decreased after treatment; metals present in the feed at concentrations lower than 3 mg/l had removals higher than 40% (e.g., Al, Fe, Mn, Sr, Zn, Ba, Ti, V, Cu). Statistically significant changes in the sludge concentration of Mn and Fe occurred after the temperature shifts in the MLSS of the Mg-SBR. The Mn concentration decreased from 2.50 to 1.48 ± 0.15 mg/g MLSS and the Fe concentration increased from 2.86 to 4.04 ± 0.95 mg/g MLSS after the temperature shift. The Mn and Fe concentrations in the other reactors remained relatively constant. The Fe concentration in the SRT-SBR also tended to increase after the shift, but this change from 3.17 to 3.76 ± 0.58 mg/g MLSS was not statistically significant. The K concentration decreased from 2.4 to 1.4 ± 0.5 mg/g MLSS and the Ca concentration increased from 17.5 to 27.8 ±5.1 mg/g MLSS in the SRT-SBR after the temperature shift. The K concentrations remained relatively constant in the other reactors. The Na concentration in the solids was highly variable in all the SBRs, which was reflected by large standard deviations, except in the Mg-SBR, where the Na concentration, although variable, averaged 6.3 ± 2.8 mg/g MLSS. The Na concentration ranged from 0–0.4 to 15.8–23.4 mg/g MLSS in the other reactors. The remaining metal concentrations in the sludge solids from the reference SBR, Mg-SBR, and SRT-SBR remained unchanged after the temperature shifts from 30 to 45 ◦ C. All the metal concentrations in the sludge from the control SBR remained relatively constant during operation at a constant temperature of 30 ◦ C. Higher concentrations of Al, Mn, Ba, Fe, Zn, and Ti were measured in the treated effluent after the temperature shift than before the shift. These metal concentrations were associated with higher ESS levels, and these metals were detected in the solids as well. 3.3. Sludge dissociation constants at 30 ◦ C before the temperature shift The sludge dissociation constants before the temperature shift, at a constant temperature of 30 ◦ C, tended to
decrease with increasing CaCl2 solution concentrations (0.0005, 0.005, and 0.5 M) in the SBRs at SRT∼20 days (reference and control SBRs) and the SRT-SBR (Fig. 3). The decrease in the dissociation constants with increasing CaCl2 concentration from 0.0005 M to 0.5 M before the shift was statistically significant (p = 0.04) in the reference SBR. The reproducibility in sludge floc characteristics at 30 ◦ C is indicated by the similar values of dissociation constants measured in the sludge from the reference SBR before the shift and the control-SBR operated constantly at 30 ◦ C. Different concentrations of KCl solutions (0.0005, 0.005, and 0.3 M) and urea solutions (2 and 8 M) had no significant impact on the sludge dissociation constants (Figs. 4 and 5) of the four SBRs before the temperature shift. Nevertheless, the dissociation constants tended to increase at the highest urea concentration of 8 M, and urea caused the highest sludge dissociation constants from the four compounds tested. Increasing the EDTA solution concentration from 20
Fig. 3. Sludge floc dissociation constants (absorbance/g MLSS/washing) before and after the temperature shift at different CaCl2 concentrations in the four SBRs (A = 0.0005 M, B = 0.005 M, and C = 0.5 M). The statistically significant differences in the dissociation constants before and after the shift are C in the reference SBR (p = 0.0195); C in the Mg-SBR (p < 0.05); and A, B, and C in the SRT-SBR (p < 0.05).
Fig. 4. Sludge floc dissociation constants (absorbance/g MLSS/washing) before and after the temperature shift at different KCl concentrations in the four SBRs (A = 0.0005 M, B = 0.005 M, and C = 0.3 M). The statistically significant differences in the dissociation constants before and after the shift are B in the Mg-SBR (p < 0.05) and B and C in the SRT-SBR (p < 0.05).
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Fig. 5. Sludge floc dissociation constants (absorbance/g MLSS/washing) before and after the temperature shift at two different urea concentrations in the four SBRs (A = 2 M and B = 8 M).
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Fig. 7. Sludge floc dissociation constants (absorbance/g MLSS/washing) in water before and after the temperature shift in the four SBRs. The difference in the dissociation constants before and after the shift is statistically significant in the SRT-SBR (p < 0.05).
3.4. Sludge dissociation constants after the temperature shift from 30 to 45 ◦ C
Fig. 6. Sludge floc dissociation constants (absorbance/g MLSS/washing) before and after the temperature shift at two different EDTA concentrations in the four SBRs (A = 20 mg/l and B = 300 mg/l). The statistically significant differences in the dissociation constants before and after the shift are B in the Mg-SBR (p < 0.05) and B in the SRT-SBR (p < 0.05).
to 300 mg/l increased the sludge dissociation constants in the four SBRs before the shift (Fig. 6). The dissociation constants at 300 mg EDTA/l were significantly higher than those at 20 mg EDTA/l in the Mg-SBR (at the 90% confidence level; p = 0.052) and in the SRT-SBR (p = 0.003). These results are in general agreement with a previous study on sludge floc stability conducted by Liao et al. [31] at lower SRTs (4, 9, 16, and 20 days) than those studied here using the same chemical agents. Liao et al.’s [31] results also showed that lower sludge dissociation constants occur with increasing concentrations of CaCl2 . The results also agree in that higher dissociation constants are observed at 300 mg EDTA/l than at 20 mg EDTA/l, in that the dissociation constants did not decrease with increasing KCl concentration at SRTs > 20 days, and in that increasing the urea solution concentration from 2 to 8 M has no significant or little effect on the dissociation constants. In addition, the dissociation constants from the CaCl2 washings were lower than those from the KCl washings, as also reported by Liao et al. [31].
The overall sludge floc structure became less stable as a result of the temperature shift, as observed from the tendency of the dissociation constants from the reference SBR to increase after the shift with the addition of CaCl2 , KCl, and EDTA (Figs. 3, 4, and 6), and water (Fig. 7). Washings with urea solutions had no significant effect on the dissociation constants after the temperature shift in any of the SBRs. The sludge flocs from the reference SBR after the temperature shift tended to deflocculate more easily (greater dissociation constants) with washings of CaCl2 and KCl solutions than with washings of EDTA and urea solutions. The percentage increase in the values of the dissociation constants of the reference SBR after the shift with respect to the average values before the shift were from 32 to 132% for CaCl2 , from 60 to 120% for KCl, from 73 to 82% for EDTA, and from −2 to 12% for urea. The sludge at an SRT∼33 days (from the SRT-SBR) deflocculated as easily as that from the reference SBR when washed with CaCl2 after the temperature shift (68–188% increase in dissociation constants in the SRT-SBR, Fig. 3). However, the SRT-SBR flocs were more stable after the shift when washed with KCl and EDTA solutions, since the KCl and EDTA dissociation constants in the SRT-SBR after the shift were similar to or not much higher than those before the shift. In addition, the increase in the KCl and EDTA dissociation constants in the SRT-SBR after the shift was smaller (39–76% for KCl and 37–47% for EDTA) than that in the reference SBR (60–120% for KCl and 73–82% for EDTA). The sludge conditioned with Mg (from Mg-SBR) showed the highest stability with respect to the reference SBR sludge and the sludge at SRT∼33 days when treated with the four chemical agents (Figs. 3–6) and water (Fig. 7). The lowest dissociation constants and the lowest increments in dissociation constants with respect to the reference SBR and SRT-
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SBR were measured with the sludge from the Mg-SBR after the temperature shift. In addition, sludge magnesium enrichment helped maintain floc stability after the temperature shift. A comparison of the sludge flocs from the four SBRs under light microscopy showed that the sludge flocs predominating in the Mg-SBR and the SRT-SBR were denser with less diffuse edges than the flocs predominating in the reference and control SBRs.
4. Discussion 4.1. Sources of a more negatively charged sludge The temperature shifts from 30 to 45 ◦ C consistently caused a net increase in the negative sludge surface charge (Fig. 2). This suggests that the sludge also became less hydrophobic, and became more negative, based on the work of Liao et al. [22], who found a significant positive correlation between sludge hydrophobicity and surface charge. The causes of a more negative sludge charge due to the temperature shift can only be hypothesized. The adsorption of extracellular polymeric substances (EPS) released from the floc matrix and of lyzed soluble microbial products bearing negatively charged groups onto remaining sludge flocs could have rendered the sludge more negatively charged. More anionic compounds on the flocs could also be exposed by the release of EPS. Negative sludge surface charge under neutral conditions has been attributed to the presence of anionic functional groups (e.g., sulfate, hydroxyl, and carboxyl groups) on the sludge floc surface [35,36]. A more negative surface charge in sludge particles has been observed after the release of calcium and organics into sludge supernatant from floc EPS due to dilution with deionized water [2]. Supporting evidence that this adsorption process could take place was obtained when fresh sludge was exposed to the supernatant from sludge previously subjected to a temperature shift, and the sludge surface charge, originally at −0.234 ± 0.026 meq/g MLSS, became more negative, up to −0.268 ± 0.045 meq/g MLSS, 5 min after supernatant addition, and up to −0.314 ± 0.033 meq/g MLSS, 1.5 h after supernatant addition. EPS biopolymers with anionic functional groups tend to increase the negative sludge surface charge, as experimentally shown by Mikkelsen et al. [37] using an anionic detergent. This adsorption process could be collateral to a major cause driving the more negatively charged sludge. The increase in the negative sludge surface charge is not an immediate response and manifests itself within 1–2 to 8 days (Fig. 2). Floc fragmentation would also increase the sludge surface area per g of sludge, thereby increasing the negative sludge surface charge per mass of sludge [33]. Floc fragmentation exposing more negatively charged surface explains the extra consumption of cationic polymer once charge neutralization has been achieved and sludge fragmentation occurs
due to increased shear stress [38]. In this case, increased sludge surface charge is a consequence of deflocculation rather than a cause of it. The more negatively charged sludge could also be influenced by the proliferation of more negatively charged filaments. The impact of the types of filamentous organisms on sludge physicochemical properties is poorly understood. However, a decrease in sludge hydrophobicity has been correlated to lower abundance of Microthrix parvicella with increasing temperature from 5 to 20 ◦ C [39]. The increase in negative sludge surface charge may arise from bacterial physiological stress responses. The degree of unsaturation of fatty acids in the cytoplasmic membrane is thought to control the cellular stress response to temperature changes [40]. Heat shocks could increase cytoplasmic membrane flexibility, which could trigger microorganisms to counteract this by increasing the levels of saturated fatty acids in the membrane phospholipids in order to decrease membrane fluidity. Saturation of phospholipids could be achieved by negatively charged groups that render the cell surface more negatively charged. An increase in saturated fatty acids in the cytoplasmic membrane as a primary heat shock sensor has been correlated to the heat shock induction of mRNA transcripts in yeasts [41]. Gaughran [42] reported experiments showing the occurrence of more saturated fatty acids in bacteria (Bacillus subtilis) and moulds (Aspergillus niger) at higher temperatures (e.g., from 14 to 38 ◦ C), even above optimal-growth temperatures. A decrease in cytoplasmic membrane fluidity, transforming the liquid crystalline membrane into a gel-phase state, results from cold shocks and induces an increase in membrane unsaturated fatty acids that lower the membrane phospholipid melting point and add membrane flexibility [40]. In addition, the so-called “compatible solutes” (e.g., sulfotrehalose, diglycerolphosphate), known to be synthesized or taken up from the environment due to environmental stresses [43], could accumulate in sludge flocs when released from some cells under heat stress and promote more negatively charged flocs. 4.2. Poor floc stability and deflocculation Sludge surface charge is considered to influence sludge floc structural stability and floc formation due to the interaction of electrostatic forces at the solid–liquid interface of sludge particles [3,37]. A more negatively charged sludge (i.e., less hydrophobic) has been correlated with increased ESS [22] and with increased floc shear sensitivity [37]. This agrees with the correlation between a more negative sludge charge and increased ESS levels in our studies after the temperature shifts [44]. More negatively charged and less hydrophobic sludge floc surfaces lead to electrostatic repulsion among floc particles and floc affinity to remain in suspension in water, promoting weaker floc structures and decreasing bioflocculation. Lower sludge hydrophobicity and presumably more negatively charged sludge have been associated with deflocculating sludge under phenol disturbances and is
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partially explained by the effect of cellular components released by lysis [8]. The increase in sludge hydrophobicity, in contrast, is known to foster cell adhesion and better bioflocculation [20,22,45] and granulation [30]. How hydrophobicity determines negative surface charge or vice versa has not been determined. The repulsive electrostatic double-layer forces, as per the Derjaguin–Landau–Verwey–Overbeek (DLVO) theory, may further induce the process of deflocculation by temperature shifts. More negatively charged sludge surfaces enhance electrostatic double-layer repulsion among particles due to increased surface charge density. A loss of floc structure stability occurred due to the increase in temperature, as indicated by greater dissociation constants (CaCl2 , KCl, EDTA, and water) after the temperature shift from 30 to 45 ◦ C in the reference SBR. The temperature shifts appear to render the sludge flocs more susceptible to changes in ionic strength than to chelating by EDTA or to the disruption of protein hydrogen bonds by the chaotropic effect of urea. This difference in effects is demonstrated by larger deflocculation (greater dissociation constants after than before the shift) occurring with washings of CaCl2 and KCl solutions than with washings of EDTA and urea solutions. In addition, the high CaCl2 dissociation constants (Fig. 3) with high (0.5 M) and low (0.0005 M) CaCl2 concentrations with respect to the constants with a medium concentration (0.005 M) suggests that the temperature shift modified the sludge flocculation-controlling mechanisms. Zita and Hermansson [3] also reported this behavior of CaCl2 dissociation constants with activated sludge flocs. Although high dissociation constants at low ionic strength and low dissociation constants at medium ionic strength could be explained by the DLVO theory, the increase in dissociation constants at high ionic strength remained unexplained. It is possible that the addition of a medium concentration of Ca2+ (0.005 M) to the temperature-weakened flocs promoted flocculation of dispersed material in this case per the DLVO theory. Nevertheless, the disturbance of sludge flocs by addition of CaCl2 during the stability tests cannot be explained based on the DLVO theory, similarly to the addition of sodium by Higgins and Novak [25]. Probably an increase in Ca concentration decreases flocculation in a similar way, as high Ca concentrations have been observed to decrease the complexation of Ca with dissolved organic matter in water via carboxyl sites [47]; this mechanism, however, has not been clearly established. The addition of K+ could also increase the ratio of monovalent to divalent cations (>2), disturbing divalent cationic bridging via ionexchange [25,48]; however, the role of K+ in deflocculation is not completely understood either. EPS solubilization may have further weakened the structure of the remaining flocs, as observed during anaerobic storage when EPS degradation weakened sludge floc structure and impaired dewaterability [46]. The occurrence of sludge deflocculation under temperature shifts from 30 to 45 ◦ C was determined in batch ex-
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periments, as reported elsewhere [10]. However, further evidence of sludge deflocculation contributing to ESS levels was obtained when larger sludge dissociation constants were measured under different chemical agents after the temperature shift (Figs. 3, 4, and 6). This indicated weaker and physically less stable sludge flocs after the temperature shift from which EPS was solubilized and small floc particles could have been eroded. The increase in the treated effluent concentrations of Al, Mn, Ba, Fe, Zn, and Ti, known to be associated with the MLSS, identifies the origin of suspended solids in the effluent as deflocculated sludge. In addition to the more negative sludge surface charge, the temperature shifts appear to have led to the release of manganese into solution and to the microbial excretion of iron either as Fe(II) or as Fe(III). This behavior of the Mn and Fe species could not be explained; however, Fe(III) is known to aid in flocculation via EPS bridging [14]. It is probable that Fe(III) predominated in the sludge, forming bonds with EPS, or was entrapped in the floc matrix as Fe(OH)3 , as observed in activated sludge from a nutrient removal plant [49]. The role of metals in bioflocculation and deflocculation remains a research area in need of more investigation. 4.3. Impact of sludge manipulation on floc stability Sludge magnesium enrichment was successful in promoting structurally stronger flocs with less negative surface charge. However, sludge deflocculation during the temperature shift still occurred, as assessed by increased ESS levels, decreased SCOD removals, and increased turbidity [44]. The relatively unaffected sludge surface charge and floc structural strength after the temperature shift in the Mg-SBR suggest that deflocculation of Mg-enriched sludge was not related to any drastic change in these sludge physicochemical properties. Charge neutralization of negatively charged groups in the sludge by magnesium ions explains the less negatively charged sludge in the Mg-SBR before and after the shift in comparison to the other reactors, and Mg accumulation in the sludge during early acclimation. Mg2+ has been reported to decrease the negative sludge surface charge [19,25]. Magnesium ions, like other divalent cations (e.g., Ca2+ ), are considered to increase sludge floc stability via divalent cationic bridging [24,25], whereby negatively charged functional groups in the EPS are joined by cations forming polymeric bridges among bacteria and/or floc particles. These divalent cations interact with negative radical groups, such as carboxylic, hydroxyl, and phosphate groups, available in the sludge EPS and bacterial cell surfaces. In a more mechanistic sense, Mg2+ has been proposed to preferentially bind to DNA [19] and to stabilize the outer cell membrane, resulting in increased solvent tolerance [50]. Mg2+ may also play a more selective role in bridging due to its smaller size than Ca2+ , or Mg2+ may stabilize the lectin-like structure or act
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as binding sites for the activity of lectin-like proteins, known to be involved in cell adhesion and aggregation [13]. Similar sludge dissociation constants from the Mg-SBR before and after the temperature shift with CaCl2 and KCl solutions (Figs. 3 and 4) support the ideas that Mg2+ promotes polymeric bridging, that Ca2+ and K+ are unable to form more polymeric bridges, and that K+ is unable to decrease particle repulsion and favor flocculation, as suggested by the DLVO theory. Mg2+ cationic bridging is not significantly affected by the temperature shift, but this is not the governing mechanism maintaining floc stability, since deflocculation still occurred after the temperature shift. The sludge flocs at a higher SRT of approximately 33 days were less negatively charged than the flocs from the reference and control SBRs (Fig. 2), but had sludge surface charge similar to that on the flocs from the Mg-SBR. The 33d-SRT flocs were also as strong and compact as the Mg-SBR flocs. This less negative sludge surface charge and the visually denser structure of the flocs confirm Liao et al.’s [31] results in that higher SRTs induce the formation of compact and strong sludge flocs. Experimentally determined sludge-dispersion coefficients from bioflocculation kinetics have indicated that the degree of sludge dispersion varies inversely with the SRT and that better bioflocculation is induced at higher SRT and lower HRT [51]. Nevertheless, the 33d-SRT sludge flocs became weaker after the temperature shift, which suggests a structural difference between the Mgenriched sludge and the 33d-SRT sludge. This structural difference could be related to the dominance of different forces and interactions in the sludge flocs, e.g., EPS enmeshment in the 33d-SRT sludge versus cationic bridging in the Mgenriched sludge. A dense EPS layer covering much of the surface of flocs and less hydrated flocs have been observed in sludge at high SRTs of 16 and 20 days [52]. The tighter structure from the Mg-enriched flocs and the flocs at an SRT of approximately 33 days could have prevented the penetration of urea and EDTA, acting as a physical barrier against floc disruption before and after the shift due to their chaotropic and chelating effects, respectively. This would explain why the sludge dissociation constants from the Mg-SBR and SRTSBR were not significantly affected by the washings with urea and EDTA solutions. Sludge flocs at SRTs of 16 and 20 days have been reported to be only slightly disturbed by EDTA and urea solutions due to a compact floc structure and a compact layer of EPS encapsulating the flocs, probably with high lipid content [31].
5. Conclusions Activated sludge floc structural stability and surface properties were studied under temperature shifts from 30 to 45 ◦ C. Floc structural stability was assessed by subjecting sludge flocs to different chemical agents in suspending solu-
tions, and sludge surface charge was measured before and after the temperature shift. The main conclusions drawn from this work are: (1) The temperature shift from 30 to 45 ◦ C deteriorates the flocculating physicochemical properties of the sludge, as assessed by the consistent increase in the negative surface charge of activated sludge flocs. (2) Magnesium sludge enrichment and a high sludge retention time (SRT) of 33 days subdued the increase in the negative sludge surface charge in comparison with the increase in the negative sludge surface charge in the nonmanipulated reactor and promoted a less negatively charged sludge. Magnesium ions appear to neutralize the negatively charged groups on the sludge flocs, which supports the idea that magnesium ions foster bioflocculation via polymeric bridging. (3) Magnesium enrichment of the sludge and a higher SRT of 33 days promoted structurally stronger flocs, as assessed by lower dissociation constants under different chemical agents. Nevertheless, the more compact and stronger floc structure obtained by magnesium enrichment and an SRT of 33 days were unsuccessful in significantly decreasing sludge deflocculation under the temperature shift from 30 to 45 ◦ C. Particularly, the 33dSRT sludge flocs became weaker than the Mg-enriched flocs after the temperature shift, suggesting a structural difference between the Mg-enriched and the 33d-SRT sludge. (4) Manganese and iron appear to be released by the sludge under the temperature shift; however, this behavior remains unexplained. Further research is required to clarify the role of metal species in deflocculation and the origin of the increase in negative sludge surface charge resulting from the temperature shift from 30 to 45 ◦ C.
Acknowledgments The authors acknowledge financial support from the members of the Consortium “Minimizing the Impact of Pulp and Paper Mill Discharges” at the Pulp & Paper Centre, University of Toronto: Aracruz Celulose S.A., Carter Holt Harvey Tasman, Domtar Inc., Eka Chemicals Inc., GeorgiaPacific Corporation, Irving Pulp and Paper Limited, Japan Carlit Co. Ltd., ERCO Worldwide (formerly Sterling Pulp Chemicals, Ltd.), and Tembec Inc. In addition, financial support from the Government of Ontario/Du Pont Graduate Scholarship in Science and Technology is gratefully acknowledged, as well as partial support from the Natural Sciences and Engineering Research Council (NSERC) of Canada. The authors thank Amy Lo and Doug Craig at Domtar Inc. for help in obtaining mill effluent samples.
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[28] B.-M. Wilén, B. Jin, P. Lant, Water Res. 37 (9) (2003) 2127. [29] B.Q. Liao, Physicochemical Studies of Microbial Flocs, Ph.D. thesis, University of Toronto, 2000. [30] J.-H. Tay, Q.-S. Liu, Y. Liu, Water Sci. Technol. 46 (4–5) (2002) 13. [31] B.Q. Liao, D.G. Allen, G.G. Leppard, I.G. Droppo, S.N. Liss, J. Colloid Interface Sci. 249 (2002) 372. [32] J.W. Morgan, C.F. Forster, L. Evison, Water Res. 24 (6) (1990) 743. [33] L.H. Mikkelsen, Water Res. 37 (10) (2003) 2458. [34] APHA, Standard Methods for the Examination of Water and Wastewater, 18th ed., American Public Health Association, Washington, DC, 1992. [35] C.F. Forster, Water Res. 2 (1968) 767. [36] C.F. Forster, Water Res. 5 (1971) 861. [37] L.H. Mikkelsen, A.K. Gotfredsen, M.L. Agerbæk, P.H. Nielsen, K. Keiding, Water Sci. Technol. 34 (3–4) (1996) 449. [38] S.K. Dentel, M.M. Abu-Orf, C.A. Walker, Chem. Eng. J. 80 (2000) 65–72. [39] S. Knoop, S. Kunst, Water Sci. Technol. 37 (4–5) (1998) 27. [40] S. Phadtare, K. Yamanaka, M. Inouye, in: G.S.A.R. Hengge-Aronis (Ed.), Bacterial Stress Responses, ASM Press, Washington, DC, 2000, pp. 33–45. [41] L. Carratù, S. Franceschelli, C.L. Pardini, G.S. Kobayashi, I. Horvath, L. Vigh, B. Maresca, Proc. Natl. Acad. Sci. USA 93 (Apr. 1996) 3870. [42] E.R.L. Gaughran, Bacter. Rev. 11 (1947) 189. [43] E. Bremer, R. Krämer, in: G.S.A.R. Hengge-Aronis (Ed.), Bacterial Stress Responses, ASM Press, Washington, DC, 2000, pp. 79–97. [44] F. Morgan-Sagastume, D.G. Allen, Water Sci. Technol., in press. [45] A.-C. Olofsson, A. Zita, M. Hermansson, Microbiology 144 (1998) 519. [46] P.H. Nielsen, B. Frølund, K. Keiding, Appl. Microbiol. Biotechnol. 44 (1996) 823. [47] Y. Lu, H.E. Allen, Water Res. 36 (2002) 5083. [48] S.N. Murthy, J.T. Novak, Water Sci. Technol. 37 (4–5) (1998) 317. [49] H. Rasmussen, P.H. Nielsen, Water Res. 30 (3) (1996) 551. [50] J. Kieboom, J.A.M. de Bont, in: G.S.A.R. Hengge-Aronis (Ed.), Bacterial Stress Responses, ASM Press, Washington, DC, 2000, pp. 393– 402. [51] G.A. Ekama, J.L. Barnard, F.W. Gunther, P. Krebs, J.A. McCorquodale, D.S. Parker, E.J. Wahlberg, Secondary Settling Tanks: Theory, Modelling and Operation, in: Scientific and Technical Report Series, vol. 6, IWA Publishing, London, 1997. [52] S.N. Liss, B.Q. Liao, I.G. Droppo, D.G. Allen, G.G. Leppard, Water Sci. Technol. 46 (1–2) (2002) 431.
Physicochemical properties and stability of activated sludge flocs under temperature upshifts from 30 to 45 ◦ C F. Morgan-Sagastume, D.G. Allen ∗ Department of Chemical Engineering and Applied Chemistry, Pulp & Paper Centre, University of Toronto, 200 College Street, Toronto, Ontario M5S 3E5, Canada Received 21 March 2004; accepted 6 August 2004 Available online 21 September 2004
Abstract The impacts of temperature shifts from 30 to 45 ◦ C on the structural stability and surface charge of activated sludge flocs were assessed in four sequencing batch reactors (SBRs) treating pulp and paper mill effluent. The improvement in floc stability was tested by sludge magnesium enrichment in one SBR and by operating another reactor at a high sludge retention time (SRT) of 33 days. Floc stability was characterized by dissociation constants with solutions of CaCl2 , KCl, urea, and ethylenediamine tetraacetate (EDTA). Surface charge was assessed by cationic–anionic titration and metals concentrations were also determined. The temperature shift consistently caused an increase in the negative sludge surface charge from approximately −0.180 to −0.300 meq/g MLSS. Magnesium enrichment and a high SRT of 33 days promoted less negatively charged sludge, dampened the increase in negative sludge surface charge, and yielded structurally stronger flocs; however, sludge deflocculation still occurred. Manganese and iron appeared to be released by sludge under the temperature shift. It was concluded that the temperature shift deteriorates the flocculating physicochemical properties of the sludge and that better floc stability achieved by magnesium enrichment and a high SRT is not enough to stop deflocculation. Further research is required to clarify the origin of the increase in negative sludge surface charge, the role of metals, and the governing factors in sludge deflocculation under such temperature shifts. 2004 Elsevier Inc. All rights reserved. Keywords: Activated sludge; Floc stability; Surface charge; Temperature upshifts; Metals; Deflocculation; Dissociation constants
1. Introduction The capacity of sludge biomass to remain flocculated, to settle fast, and to compact determines the efficiency of solid– liquid separation in the activated sludge process. Although good sludge settling and compaction result in the separation of sludge biomass from treated effluent, poor sludge flocculation can cause increased levels of effluent suspended solids (ESS). Poor sludge flocculation is associated with weak floc structure of part of the biomass, with pinpoint flocs and with free-swimming cells in suspension. All these conditions lead to solids discharges, to the loss of biocatalyst used in pollutant degradation, and to the discharge of bioactive compounds adsorbed onto biomass. * Corresponding author. Fax: +1-416-971-2106.
E-mail address: [email protected] (D.G. Allen). 0021-9797/$ – see front matter 2004 Elsevier Inc. All rights reserved. doi:10.1016/j.jcis.2004.08.026
Weakening of bioaggregate structure and deflocculation appear to be common responses of suspended sludge biomass to transient operating conditions and environmental stresses, such as shifts in substrate loadings [1], in ionic strength [2,3], and in the concentration of dissolved oxygen [4–6] and toxic compounds [7,8] or metals [9]. Activated sludge deflocculation has also been shown to occur due to mesophilic–thermophilic temperature transients from 30–35 to 45 ◦ C [10]. Temperature shifts from 30–35 to 45 ◦ C in treatment plants of pulp and paper mills result from seasonal variations and the operation of batch units and shutdowns/start-ups in the upstream mill. Sludge deflocculation under such temperature shifts impairs treatment performance by increasing ESS levels and decreasing the removal efficiency of chemical oxygen demand (COD), as reported for sequencing batch reactors (SBRs) treating bleached kraft pulp and paper mill effluent [11].
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A means of preventing sludge deflocculation from happening is increasing the structural stability of sludge flocs. The factors determining sludge bioflocculation, and thus floc structure and stability, have been studied in the past few decades, as recently reviewed by Liss [12]. Sludge floc structure and stability are influenced by physicochemical and microbial factors. Physicochemical factors include divalent cationic bridging of EPS [13]; Fe(III) polymer bridging or flocculation via Fe(OH)3 [14,15]; polymer bridging, originally proposed by Pavoni et al., 1972 (see [13,16,17]); electrostatic double-layer repulsive and van der Waals attractive interactions [3]; alginate gel formation with Ca2+ [18]; hydrophobic interactions [19]; and steric interactions [20]. Microbial factors include relative amount, of filamentous organisms [21] and sludge age [22]. Due to the varying nature of sludge, one or several of these factors may play a predominant role in floc stability in a given type of sludge; however, quantifying these relative forces or interactions in sludge flocs is challenging and not yet feasible [23]. Improving sludge bioflocculation characteristics through the addition of cations can increase sludge settleability and also floc structural stability. Improved sludge settling characteristics (assessed by low sludge volume indices (SVIs)) and sludge dewatering (assessed by low capillary suction time (CST) and specific resistance to filtration (SRF)) can be achieved by adding divalent cations (Ca2+ and Mg2+ ) to the sludge. These divalent cations are considered to improve bioflocculation by cationic bridging [24]. Particularly important in improving sludge settling and dewatering are 2+ the ratio of monovalent (Na+ , K+ , NH+ 4 ) to divalent (Ca 2+ 2+ 2+ and Mg ) cations and the ratio of Ca to Mg in the feed. Ratios of monovalent to divalent cations greater than 2 and Ca2+ /Mg2+ ratios significantly different from 1 have been related to deteriorated floc properties [25,26]. Increased sludge Mg concentrations have been correlated with better sludge settleability and compressibility [27]. The replacement of Ca2+ in sludge by ion exchange with counterions such as H+ , Na+ , and K+ was observed to increase the specific resistance to filtration [18]. Sludge bioflocculation is governed by sludge surface properties, such as hydrophobicity, surface charge, and the composition of extractable extracellular polymeric substances (EPS) and of total polymers in the sludge [28]. A higher ratio of proteins to carbohydrates in EPS rather than the total amount of EPS and a less negatively charged and more hydrophobic sludge have been correlated with lower ESS and therefore to better sludge flocculation [22,29]. Also, increased cell hydrophobicity has been correlated with floc and granule formation; cell hydrophobicity is known to lead to bacterial adhesion [30]. A way of inducing a less negatively charged and more hydrophobic sludge, and therefore structurally stronger sludge flocs, would be by operating at higher SRTs (e.g., greater than 12 days), as demonstrated by Liao et al. [22,29]. Insights into the mechanisms controlling the structural stability of activated sludge flocs can be obtained by investi-
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gating the response of sludge flocs subjected to different environmental conditions. For example, Liao et al. [31] tested the effects of electrostatic interactions, ionic interactions, and the absence of hydrogen bonds on sludge floc stability. Activated sludge flocs were exposed to suspending solutions of varying ionic strength and cationic valence and with varying concentrations of ethylenediamine tetraacetate (EDTA) and urea. The goal of this work was to assess the changes in surface properties and structural stability of manipulated and nonmanipulated activated sludge flocs under temperature shifts from 30 to 45 ◦ C in sequencing batch reactors (SBRs) treating bleached kraft pulp mill effluent. Sludge cationic enrichment and high sludge ages of 33 days were assessed as ways to increase floc structural stability in activated sludge. Floc structural stability was evaluated by subjecting the sludge flocs to different suspending solutions with various ionic strength, cationic valence, and urea and EDTA concentration.
2. Experimental procedures 2.1. Sequencing batch reactors (SBRs) Four parallel sequencing batch reactors (SBRs) were operated to treat primary clarified bleached kraft pulp mill effluent for 80 days. Two operating strategies for improving floc structural stability were assessed: sludge cationic enrichment with Mg2+ and a high sludge retention time (SRT) of 33 days. Each strategy was tested independently in two SBRs (Mg-SBR and SRT-SBR). The four SBRs were operated in three 8-h cycles per day, with a reaction volume of 1.55 l and a hydraulic retention time of 14 h. Each 8-h cycle consisted of 25 min anoxic filling with mixing, 385 min reaction with mixing and aeration, 60 min settling, and 10 min discharge. The dissolved oxygen levels were above 2–3 mg/l and the pH was between 7.4 and 8.4 in the four SBRs. The raw mill effluent was transported at 4 ◦ C to our laboratory immediately after collection and then frozen at −20 ◦C. The wastewater was thawed as required (about 84 l/week) and conditioned with NH4 Cl (Mallinckrodt Inc., Paris, KY) and (NH4 )2 HPO4 (BDH Inc., Toronto, ON) to achieve a solubleCOD:N:P ratio of 200:5:1. The soluble chemical oxygen demand (SCOD) concentration in the feed was relatively constant in the four SBRs (reference SBR: 282 ± 38 mg/l; Mg-SBR: 284 ± 35 mg/l; SRT-SBR: 296 ± 39 mg/l; control SBR: 283 ± 52 mg/l) although it decreased by less than 100 mg/l over the time of operation due to storage. The biomass inoculum (0.5 l/reactor) was return activated sludge from the same treatment plant from where the wastewater was collected. Details of the experimental setup are reported elsewhere [10,11]. Sludge cationic enrichment was achieved by adding MgCl2 ·6H2 O (Caledon Laboratories Ltd., Georgetown, ON) to the prepared feed of one SBR (Mg-SBR) in order to
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Table 1 Similarities and differences in operating conditions among the four parallel SBRs Characteristic
Reference SBR
Mg-SBR
SRT-SBR
Control-SBR
Sludge retention time (SRT) (d) Temperature shift from 30 to 45 ◦ C Sludge manipulation
∼20 YES None
∼20 YES Mg enrichment
∼33 YES Higher SRT
∼20 NO, T = 30 ◦ C None
achieve an equivalent Ca2+ /Mg2+ ratio close to 1 (0.9 ± 0.1 as Ca/Mg) and an equivalent ratio of monovalent (Na+ ) to divalent (Ca2+ and Mg2+ ) cations close to 2 [2.1 ± 0.5 as Na/(Ca and Mg)]. The total Mg concentration in the prepared feed was increased from 0.9 to 3.7 ± 0.5 meq/l. The Ca concentration in the raw mill effluent was 3.4 ± 0.4 meq/l. Ninety-eight percent of the total of K and Na concentrations in the mill effluent was Na (288 ± 39 mg Na/l). One of the four SBRs was operated at an SRT of 33 days (SRT − SBR = 33 ± 9 days), and the other three SBRs were operated at an SRT of ∼20 days (control SBR: 20 ± 6 mg/l; Mg-SBR: 20 ± 7 days; reference SBR: 22 ± 13 days). The SRTs were maintained by wasting mixed liquor as a function of ESS and mixed liquor total suspended solids (MLSS) concentrations every 2 days. The sludge in two of the three SBRs operating at an SRT of 20 days was not manipulated, and one of these two reactors (reference-SBR) was subjected to a temperature upshift while the other one (control-SBR) was operated at a constant temperature of 30 ◦ C as a control. Table 1 summarizes the similarities and differences in operating conditions among the four SBRs. 2.2. Temperature upshifts The temperature shifts from 30 to 45 ◦ C in three of the SBRs (reference-SBR, Mg-SBR, and SRT-SBR) were conducted in sequence, as shown in Fig. 1., within 1.5 to 2.9 h of the reaction phase of an operating cycle. The sludge from the reference and control SBRs was mixed and redistributed between these two SBRs on day 30, before the first shift. The reactors’ temperature was monitored with 76-mm mercury thermometers. Deep-chamber water baths (reference, Mg-, and SRT-SBRs: 33 L, 1295 PC, VWR Scientific, Mississauga, ON; MeOH SBR: Precision Model 188, Precision Scientific Inc.) connected to water jackets surrounding each reactor were used to control the reactors’ temperatures. The accuracy of the temperature readings was ±1 ◦C. 2.3. Sludge surface charge, pH, and analysis of metals Sludge surface charge was determined by cationic– anionic titration [22,32], and reported as meq of anionic polymer/g MLSS. A 0.002-N hexadimethrine bromide (Polybrene) solution and a 0.001-N sodium-salt–polyanetholesulfonic-acid solution were used as the cationic and the anionic standards, respectively. The detailed protocol is outlined elsewhere [10]. Although a recent study [33] indicates that colloid titration tends to overestimate sludge sur-
Fig. 1. Temperature profiles in the four parallel SBRs showing sequential temperature shifts from 30 to 45 ◦ C.
face charge, this simple technique is still useful in comparing relative values of sludge surface charge, as in this study, before and after the temperature shift and among SBRs. The sludge surface charge measured in this was may reflect the total floc sludge charge rather than the sludge surface charge alone [33]. Samples of the influent, treated effluent, and mixed liquor from the reactors were sent to Maxxam Analytics Inc. (Mississauga, Ontario) for acid-extractable metals analysis. The concentration of metals in the solids corresponds to the difference between the mixed liquor and treated effluent concentrations, divided by the MLSS concentration. The metals determined were Al, Ba, Be, Bi, Ca, Cd, Cr, Co, Cu, Fe, Pb, Mg, Mn, Mo, Ni, K, Ag, Na, Sr, Ti, W, V, and Zn. The samples were acidified with HNO3 at collection, transported at 4 ◦ C, and analyzed by inductively coupled plasma mass spectroscopy (ICP-MS) based on the United States Environmental Protection Agency’s (EPA) protocol 6020. The detection limits of the analyses were 0.001 mg/l for Mn, Ba, and Be; 0.003 mg/l for Cu, Mg, Sr, Ti, V, Zn, and Cd; 0.005 mg/l for Cr, Co, and Fe; 0.006 mg/l for Mo; 0.007 mg/l for Ag; 0.01 mg/l for Ni; 0.025 for Al and Pb; 0.04 mg/l for Bi and Ca; 0.05 mg/l for W; 0.06 mg/l for Na; and 1 mg/l for K. MLSS, MLVSS, and ESS were measured based on the Standard Methods [34]. An approximate 80-ml sample of mixed liquor was taken from each reactor during the reaction phase for measuring pH using a digital pH electrode (DigiSense, Cole–Parmer Instrument Company, Niles Illinois), conductivity (Radiometer A/S, Copenhagen, Denmark) and for MLSS samples.
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2.4. Sludge floc stability Sludge floc stability was characterized by dissociation constants determined from plots of cumulative absorbance versus number of washings with different chemical solutions, as described by Zita and Hermansson [3] and Liao et al. [31]. The dissociation constants are defined as unit of absorbance/g MLSS/washing and quantify the detachment of solids from sludge flocs due to successive washings with solutions of calcium chloride (CaCl2 ; Fisher), potassium chloride (KCl; Sigma), urea (Sigma U-5128, St. Louis, Missouri), and ethylenediamine tetraacetate (EDTA, tetrasodium salt dihydrate, Fisher). Higher dissociation constants imply a weaker sludge floc structure. The dissociation constants were determined for different solution concentrations: 0.5, 0.005, and 0.0005 M for CaCl2 ; 0.3, 0.005, and 0.0005 M for KCl; 2 and 8 M for urea; and 20 and 300 mg/l for EDTA. The absorbance was measured at 450 nm, and four or five washings were conducted with each sample. 2.5. Statistical analyses The statistical significance of differences between means from the same SBR before and after a temperature shift was assessed by using paired t-tests. The statistical significance of differences between means from two different SBRs and between means for unequal numbers of observations before and after the shift from the same SBR was assessed by ttests for independent samples and unequal variances. The statistical significance of differences between single observations and means was assessed by t-tests about the population mean (µ, assumed to be equal to the single observation) when the population standard deviation (σ ) was unknown. The statistical significance was assessed at the 95% confidence level, unless indicated, and the levels of significance (p) are reported for those cases where a significant difference was found.
3. Results 3.1. Sludge surface charge The temperature shifts from 30 to 45 ◦ C caused the sludge surface to become more negatively charged (Fig. 2). Statistically significant increases in the negative sludge surface charge were measured a few days after the shifts in the MgSBR (148% or 2.5-fold) from −0.068 ± 0.040 (day 38) to −0.169 ± 0.049 meq/g MLSS (day 45), in the reference SBR (91%) from −0.176 ± 0.059 (day 32) to −0.337 ± 0.012 meq/g MLSS (day 45), and in the SRT-SBR (22%) from −0.166 ± 0.005 (day 50) to −0.202 ± 0.001 meq/g MLSS (day 58). The more negative sludge surface charge resulting from the shift from 30 to 45 ◦ C is reproducible and consistent with our previous long-term experiments in which the sludge surface charge changed from −0.15 to
Fig. 2. Sludge surface charge before and after the temperature shift from 30 to 45 ◦ C in the four parallel SBRs. The error bars correspond to standard deviations from two samples. Arrows and labels in the figure indicate times when temperatures were shifted for various SBRs.
−0.24 meq/g MLSS due to a temperature shift from 35 to 45 ◦ C [11]. The sludge surface charge in the Mg-SBR and in the SRT-SBR before and after the shift remained less negative (>−0.202 meq/g MLSS) than that of the sludge in the reference SBR after the shift (<−0.240 meq/g MLSS). The sludge surface charge in the control SBR at a constant temperature of 30 ◦ C remained relatively constant and with values similar to those of the SRT-SBR. Mg enrichment, SRT∼33 days, and constant-temperature operation (control SBR = 30 ◦ C) promoted a less negatively charged sludge in comparison to the reference SBR (Fig. 2). The mixed liquor pH remained relatively constant in the four SBRs during the 80-day period of operation (reference SBR: pH 8.0 ± 0.2; Mg-SBR: pH 8.0 ± 0.2; SRT-SBR: pH 7.8 ± 0.2; control SBR: pH 7.9 ± 0.2), and was not significantly affected by the temperature shifts. No difference in the conductivity of sludge supernatants was detected before and after the shifts in the four SBRs. The addition of MgCl2 kept the mixed liquor conductivity in the Mg-SBR higher, as expected (reference SBR 1695 ± 120 µS/cm; Mg-SBR 2005 ± 199 µS/cm; SRT-SBR 1695 ± 108 µS/cm; control SBR 1701 ± 109 µS/cm). 3.2. Sludge metal concentrations The most abundant metal from those determined in the sludge was Ca (∼17–27 mg/g MLSS), followed by Na (∼6–13 mg/g MLSS) or Al (∼7–9 mg/g MLSS), Mg (∼3–4 mg/g MLSS), Mn (∼2–3 mg/g MLSS), Fe (∼2–4 mg/g MLSS), and K (∼2 mg/g MLSS). The sludge Mg concentration was approximately two- to threefold higher in the Mg-SBR (8.95 ± 2.32 mg/g MLSS) than in the other SBRs (reference SBR: 3.99 ± 1.64 mg/g MLSS; SRT-SBR: 3.53 ± 0.38 mg/g MLSS; and control SBR: 3.16 ± 0.31 mg/g MLSS). The original Mg concentration in the sludge solids of the Mg-SBR increased from 5.7 mg/g MLSS in the inoculum sludge to 7.2 mg/g MLSS in the
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sludge solids of the mixed liquor in the first 40 days of operation, due to the addition of Mg to the reactor’s feed. Zn and Ba concentrations in the sludge ranged from 0.1 to 1 mg/g MLSS, and the concentrations of Sr, Ti, Cu, Pb, Cr, and V were below 0.1 mg/g MLSS in the four SBRs. Ag, Be, Bi, Cd, Co, Cr, Mo, Ni, and W were not detected in the mixed liquor samples. The average concentrations of Na (∼272–322 mg/l), Ca (∼68–74 mg/l), Mg (∼10 mg/l), K (∼7–9 mg/l), and Al (∼2 mg/l) in the feed to the reactors were the highest from the metals analyzed, particularly Na, with concentrations of approximately 300 mg/l. The concentrations of Zn, Ba, Ti, V, and Cu in the feed were below 0.1 mg/l. The feed metal concentrations were relatively constant during the period of operation and among reactors. Most of the concentration of metals in the influent decreased after treatment; metals present in the feed at concentrations lower than 3 mg/l had removals higher than 40% (e.g., Al, Fe, Mn, Sr, Zn, Ba, Ti, V, Cu). Statistically significant changes in the sludge concentration of Mn and Fe occurred after the temperature shifts in the MLSS of the Mg-SBR. The Mn concentration decreased from 2.50 to 1.48 ± 0.15 mg/g MLSS and the Fe concentration increased from 2.86 to 4.04 ± 0.95 mg/g MLSS after the temperature shift. The Mn and Fe concentrations in the other reactors remained relatively constant. The Fe concentration in the SRT-SBR also tended to increase after the shift, but this change from 3.17 to 3.76 ± 0.58 mg/g MLSS was not statistically significant. The K concentration decreased from 2.4 to 1.4 ± 0.5 mg/g MLSS and the Ca concentration increased from 17.5 to 27.8 ±5.1 mg/g MLSS in the SRT-SBR after the temperature shift. The K concentrations remained relatively constant in the other reactors. The Na concentration in the solids was highly variable in all the SBRs, which was reflected by large standard deviations, except in the Mg-SBR, where the Na concentration, although variable, averaged 6.3 ± 2.8 mg/g MLSS. The Na concentration ranged from 0–0.4 to 15.8–23.4 mg/g MLSS in the other reactors. The remaining metal concentrations in the sludge solids from the reference SBR, Mg-SBR, and SRT-SBR remained unchanged after the temperature shifts from 30 to 45 ◦ C. All the metal concentrations in the sludge from the control SBR remained relatively constant during operation at a constant temperature of 30 ◦ C. Higher concentrations of Al, Mn, Ba, Fe, Zn, and Ti were measured in the treated effluent after the temperature shift than before the shift. These metal concentrations were associated with higher ESS levels, and these metals were detected in the solids as well. 3.3. Sludge dissociation constants at 30 ◦ C before the temperature shift The sludge dissociation constants before the temperature shift, at a constant temperature of 30 ◦ C, tended to
decrease with increasing CaCl2 solution concentrations (0.0005, 0.005, and 0.5 M) in the SBRs at SRT∼20 days (reference and control SBRs) and the SRT-SBR (Fig. 3). The decrease in the dissociation constants with increasing CaCl2 concentration from 0.0005 M to 0.5 M before the shift was statistically significant (p = 0.04) in the reference SBR. The reproducibility in sludge floc characteristics at 30 ◦ C is indicated by the similar values of dissociation constants measured in the sludge from the reference SBR before the shift and the control-SBR operated constantly at 30 ◦ C. Different concentrations of KCl solutions (0.0005, 0.005, and 0.3 M) and urea solutions (2 and 8 M) had no significant impact on the sludge dissociation constants (Figs. 4 and 5) of the four SBRs before the temperature shift. Nevertheless, the dissociation constants tended to increase at the highest urea concentration of 8 M, and urea caused the highest sludge dissociation constants from the four compounds tested. Increasing the EDTA solution concentration from 20
Fig. 3. Sludge floc dissociation constants (absorbance/g MLSS/washing) before and after the temperature shift at different CaCl2 concentrations in the four SBRs (A = 0.0005 M, B = 0.005 M, and C = 0.5 M). The statistically significant differences in the dissociation constants before and after the shift are C in the reference SBR (p = 0.0195); C in the Mg-SBR (p < 0.05); and A, B, and C in the SRT-SBR (p < 0.05).
Fig. 4. Sludge floc dissociation constants (absorbance/g MLSS/washing) before and after the temperature shift at different KCl concentrations in the four SBRs (A = 0.0005 M, B = 0.005 M, and C = 0.3 M). The statistically significant differences in the dissociation constants before and after the shift are B in the Mg-SBR (p < 0.05) and B and C in the SRT-SBR (p < 0.05).
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Fig. 5. Sludge floc dissociation constants (absorbance/g MLSS/washing) before and after the temperature shift at two different urea concentrations in the four SBRs (A = 2 M and B = 8 M).
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Fig. 7. Sludge floc dissociation constants (absorbance/g MLSS/washing) in water before and after the temperature shift in the four SBRs. The difference in the dissociation constants before and after the shift is statistically significant in the SRT-SBR (p < 0.05).
3.4. Sludge dissociation constants after the temperature shift from 30 to 45 ◦ C
Fig. 6. Sludge floc dissociation constants (absorbance/g MLSS/washing) before and after the temperature shift at two different EDTA concentrations in the four SBRs (A = 20 mg/l and B = 300 mg/l). The statistically significant differences in the dissociation constants before and after the shift are B in the Mg-SBR (p < 0.05) and B in the SRT-SBR (p < 0.05).
to 300 mg/l increased the sludge dissociation constants in the four SBRs before the shift (Fig. 6). The dissociation constants at 300 mg EDTA/l were significantly higher than those at 20 mg EDTA/l in the Mg-SBR (at the 90% confidence level; p = 0.052) and in the SRT-SBR (p = 0.003). These results are in general agreement with a previous study on sludge floc stability conducted by Liao et al. [31] at lower SRTs (4, 9, 16, and 20 days) than those studied here using the same chemical agents. Liao et al.’s [31] results also showed that lower sludge dissociation constants occur with increasing concentrations of CaCl2 . The results also agree in that higher dissociation constants are observed at 300 mg EDTA/l than at 20 mg EDTA/l, in that the dissociation constants did not decrease with increasing KCl concentration at SRTs > 20 days, and in that increasing the urea solution concentration from 2 to 8 M has no significant or little effect on the dissociation constants. In addition, the dissociation constants from the CaCl2 washings were lower than those from the KCl washings, as also reported by Liao et al. [31].
The overall sludge floc structure became less stable as a result of the temperature shift, as observed from the tendency of the dissociation constants from the reference SBR to increase after the shift with the addition of CaCl2 , KCl, and EDTA (Figs. 3, 4, and 6), and water (Fig. 7). Washings with urea solutions had no significant effect on the dissociation constants after the temperature shift in any of the SBRs. The sludge flocs from the reference SBR after the temperature shift tended to deflocculate more easily (greater dissociation constants) with washings of CaCl2 and KCl solutions than with washings of EDTA and urea solutions. The percentage increase in the values of the dissociation constants of the reference SBR after the shift with respect to the average values before the shift were from 32 to 132% for CaCl2 , from 60 to 120% for KCl, from 73 to 82% for EDTA, and from −2 to 12% for urea. The sludge at an SRT∼33 days (from the SRT-SBR) deflocculated as easily as that from the reference SBR when washed with CaCl2 after the temperature shift (68–188% increase in dissociation constants in the SRT-SBR, Fig. 3). However, the SRT-SBR flocs were more stable after the shift when washed with KCl and EDTA solutions, since the KCl and EDTA dissociation constants in the SRT-SBR after the shift were similar to or not much higher than those before the shift. In addition, the increase in the KCl and EDTA dissociation constants in the SRT-SBR after the shift was smaller (39–76% for KCl and 37–47% for EDTA) than that in the reference SBR (60–120% for KCl and 73–82% for EDTA). The sludge conditioned with Mg (from Mg-SBR) showed the highest stability with respect to the reference SBR sludge and the sludge at SRT∼33 days when treated with the four chemical agents (Figs. 3–6) and water (Fig. 7). The lowest dissociation constants and the lowest increments in dissociation constants with respect to the reference SBR and SRT-
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SBR were measured with the sludge from the Mg-SBR after the temperature shift. In addition, sludge magnesium enrichment helped maintain floc stability after the temperature shift. A comparison of the sludge flocs from the four SBRs under light microscopy showed that the sludge flocs predominating in the Mg-SBR and the SRT-SBR were denser with less diffuse edges than the flocs predominating in the reference and control SBRs.
4. Discussion 4.1. Sources of a more negatively charged sludge The temperature shifts from 30 to 45 ◦ C consistently caused a net increase in the negative sludge surface charge (Fig. 2). This suggests that the sludge also became less hydrophobic, and became more negative, based on the work of Liao et al. [22], who found a significant positive correlation between sludge hydrophobicity and surface charge. The causes of a more negative sludge charge due to the temperature shift can only be hypothesized. The adsorption of extracellular polymeric substances (EPS) released from the floc matrix and of lyzed soluble microbial products bearing negatively charged groups onto remaining sludge flocs could have rendered the sludge more negatively charged. More anionic compounds on the flocs could also be exposed by the release of EPS. Negative sludge surface charge under neutral conditions has been attributed to the presence of anionic functional groups (e.g., sulfate, hydroxyl, and carboxyl groups) on the sludge floc surface [35,36]. A more negative surface charge in sludge particles has been observed after the release of calcium and organics into sludge supernatant from floc EPS due to dilution with deionized water [2]. Supporting evidence that this adsorption process could take place was obtained when fresh sludge was exposed to the supernatant from sludge previously subjected to a temperature shift, and the sludge surface charge, originally at −0.234 ± 0.026 meq/g MLSS, became more negative, up to −0.268 ± 0.045 meq/g MLSS, 5 min after supernatant addition, and up to −0.314 ± 0.033 meq/g MLSS, 1.5 h after supernatant addition. EPS biopolymers with anionic functional groups tend to increase the negative sludge surface charge, as experimentally shown by Mikkelsen et al. [37] using an anionic detergent. This adsorption process could be collateral to a major cause driving the more negatively charged sludge. The increase in the negative sludge surface charge is not an immediate response and manifests itself within 1–2 to 8 days (Fig. 2). Floc fragmentation would also increase the sludge surface area per g of sludge, thereby increasing the negative sludge surface charge per mass of sludge [33]. Floc fragmentation exposing more negatively charged surface explains the extra consumption of cationic polymer once charge neutralization has been achieved and sludge fragmentation occurs
due to increased shear stress [38]. In this case, increased sludge surface charge is a consequence of deflocculation rather than a cause of it. The more negatively charged sludge could also be influenced by the proliferation of more negatively charged filaments. The impact of the types of filamentous organisms on sludge physicochemical properties is poorly understood. However, a decrease in sludge hydrophobicity has been correlated to lower abundance of Microthrix parvicella with increasing temperature from 5 to 20 ◦ C [39]. The increase in negative sludge surface charge may arise from bacterial physiological stress responses. The degree of unsaturation of fatty acids in the cytoplasmic membrane is thought to control the cellular stress response to temperature changes [40]. Heat shocks could increase cytoplasmic membrane flexibility, which could trigger microorganisms to counteract this by increasing the levels of saturated fatty acids in the membrane phospholipids in order to decrease membrane fluidity. Saturation of phospholipids could be achieved by negatively charged groups that render the cell surface more negatively charged. An increase in saturated fatty acids in the cytoplasmic membrane as a primary heat shock sensor has been correlated to the heat shock induction of mRNA transcripts in yeasts [41]. Gaughran [42] reported experiments showing the occurrence of more saturated fatty acids in bacteria (Bacillus subtilis) and moulds (Aspergillus niger) at higher temperatures (e.g., from 14 to 38 ◦ C), even above optimal-growth temperatures. A decrease in cytoplasmic membrane fluidity, transforming the liquid crystalline membrane into a gel-phase state, results from cold shocks and induces an increase in membrane unsaturated fatty acids that lower the membrane phospholipid melting point and add membrane flexibility [40]. In addition, the so-called “compatible solutes” (e.g., sulfotrehalose, diglycerolphosphate), known to be synthesized or taken up from the environment due to environmental stresses [43], could accumulate in sludge flocs when released from some cells under heat stress and promote more negatively charged flocs. 4.2. Poor floc stability and deflocculation Sludge surface charge is considered to influence sludge floc structural stability and floc formation due to the interaction of electrostatic forces at the solid–liquid interface of sludge particles [3,37]. A more negatively charged sludge (i.e., less hydrophobic) has been correlated with increased ESS [22] and with increased floc shear sensitivity [37]. This agrees with the correlation between a more negative sludge charge and increased ESS levels in our studies after the temperature shifts [44]. More negatively charged and less hydrophobic sludge floc surfaces lead to electrostatic repulsion among floc particles and floc affinity to remain in suspension in water, promoting weaker floc structures and decreasing bioflocculation. Lower sludge hydrophobicity and presumably more negatively charged sludge have been associated with deflocculating sludge under phenol disturbances and is
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partially explained by the effect of cellular components released by lysis [8]. The increase in sludge hydrophobicity, in contrast, is known to foster cell adhesion and better bioflocculation [20,22,45] and granulation [30]. How hydrophobicity determines negative surface charge or vice versa has not been determined. The repulsive electrostatic double-layer forces, as per the Derjaguin–Landau–Verwey–Overbeek (DLVO) theory, may further induce the process of deflocculation by temperature shifts. More negatively charged sludge surfaces enhance electrostatic double-layer repulsion among particles due to increased surface charge density. A loss of floc structure stability occurred due to the increase in temperature, as indicated by greater dissociation constants (CaCl2 , KCl, EDTA, and water) after the temperature shift from 30 to 45 ◦ C in the reference SBR. The temperature shifts appear to render the sludge flocs more susceptible to changes in ionic strength than to chelating by EDTA or to the disruption of protein hydrogen bonds by the chaotropic effect of urea. This difference in effects is demonstrated by larger deflocculation (greater dissociation constants after than before the shift) occurring with washings of CaCl2 and KCl solutions than with washings of EDTA and urea solutions. In addition, the high CaCl2 dissociation constants (Fig. 3) with high (0.5 M) and low (0.0005 M) CaCl2 concentrations with respect to the constants with a medium concentration (0.005 M) suggests that the temperature shift modified the sludge flocculation-controlling mechanisms. Zita and Hermansson [3] also reported this behavior of CaCl2 dissociation constants with activated sludge flocs. Although high dissociation constants at low ionic strength and low dissociation constants at medium ionic strength could be explained by the DLVO theory, the increase in dissociation constants at high ionic strength remained unexplained. It is possible that the addition of a medium concentration of Ca2+ (0.005 M) to the temperature-weakened flocs promoted flocculation of dispersed material in this case per the DLVO theory. Nevertheless, the disturbance of sludge flocs by addition of CaCl2 during the stability tests cannot be explained based on the DLVO theory, similarly to the addition of sodium by Higgins and Novak [25]. Probably an increase in Ca concentration decreases flocculation in a similar way, as high Ca concentrations have been observed to decrease the complexation of Ca with dissolved organic matter in water via carboxyl sites [47]; this mechanism, however, has not been clearly established. The addition of K+ could also increase the ratio of monovalent to divalent cations (>2), disturbing divalent cationic bridging via ionexchange [25,48]; however, the role of K+ in deflocculation is not completely understood either. EPS solubilization may have further weakened the structure of the remaining flocs, as observed during anaerobic storage when EPS degradation weakened sludge floc structure and impaired dewaterability [46]. The occurrence of sludge deflocculation under temperature shifts from 30 to 45 ◦ C was determined in batch ex-
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periments, as reported elsewhere [10]. However, further evidence of sludge deflocculation contributing to ESS levels was obtained when larger sludge dissociation constants were measured under different chemical agents after the temperature shift (Figs. 3, 4, and 6). This indicated weaker and physically less stable sludge flocs after the temperature shift from which EPS was solubilized and small floc particles could have been eroded. The increase in the treated effluent concentrations of Al, Mn, Ba, Fe, Zn, and Ti, known to be associated with the MLSS, identifies the origin of suspended solids in the effluent as deflocculated sludge. In addition to the more negative sludge surface charge, the temperature shifts appear to have led to the release of manganese into solution and to the microbial excretion of iron either as Fe(II) or as Fe(III). This behavior of the Mn and Fe species could not be explained; however, Fe(III) is known to aid in flocculation via EPS bridging [14]. It is probable that Fe(III) predominated in the sludge, forming bonds with EPS, or was entrapped in the floc matrix as Fe(OH)3 , as observed in activated sludge from a nutrient removal plant [49]. The role of metals in bioflocculation and deflocculation remains a research area in need of more investigation. 4.3. Impact of sludge manipulation on floc stability Sludge magnesium enrichment was successful in promoting structurally stronger flocs with less negative surface charge. However, sludge deflocculation during the temperature shift still occurred, as assessed by increased ESS levels, decreased SCOD removals, and increased turbidity [44]. The relatively unaffected sludge surface charge and floc structural strength after the temperature shift in the Mg-SBR suggest that deflocculation of Mg-enriched sludge was not related to any drastic change in these sludge physicochemical properties. Charge neutralization of negatively charged groups in the sludge by magnesium ions explains the less negatively charged sludge in the Mg-SBR before and after the shift in comparison to the other reactors, and Mg accumulation in the sludge during early acclimation. Mg2+ has been reported to decrease the negative sludge surface charge [19,25]. Magnesium ions, like other divalent cations (e.g., Ca2+ ), are considered to increase sludge floc stability via divalent cationic bridging [24,25], whereby negatively charged functional groups in the EPS are joined by cations forming polymeric bridges among bacteria and/or floc particles. These divalent cations interact with negative radical groups, such as carboxylic, hydroxyl, and phosphate groups, available in the sludge EPS and bacterial cell surfaces. In a more mechanistic sense, Mg2+ has been proposed to preferentially bind to DNA [19] and to stabilize the outer cell membrane, resulting in increased solvent tolerance [50]. Mg2+ may also play a more selective role in bridging due to its smaller size than Ca2+ , or Mg2+ may stabilize the lectin-like structure or act
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as binding sites for the activity of lectin-like proteins, known to be involved in cell adhesion and aggregation [13]. Similar sludge dissociation constants from the Mg-SBR before and after the temperature shift with CaCl2 and KCl solutions (Figs. 3 and 4) support the ideas that Mg2+ promotes polymeric bridging, that Ca2+ and K+ are unable to form more polymeric bridges, and that K+ is unable to decrease particle repulsion and favor flocculation, as suggested by the DLVO theory. Mg2+ cationic bridging is not significantly affected by the temperature shift, but this is not the governing mechanism maintaining floc stability, since deflocculation still occurred after the temperature shift. The sludge flocs at a higher SRT of approximately 33 days were less negatively charged than the flocs from the reference and control SBRs (Fig. 2), but had sludge surface charge similar to that on the flocs from the Mg-SBR. The 33d-SRT flocs were also as strong and compact as the Mg-SBR flocs. This less negative sludge surface charge and the visually denser structure of the flocs confirm Liao et al.’s [31] results in that higher SRTs induce the formation of compact and strong sludge flocs. Experimentally determined sludge-dispersion coefficients from bioflocculation kinetics have indicated that the degree of sludge dispersion varies inversely with the SRT and that better bioflocculation is induced at higher SRT and lower HRT [51]. Nevertheless, the 33d-SRT sludge flocs became weaker after the temperature shift, which suggests a structural difference between the Mgenriched sludge and the 33d-SRT sludge. This structural difference could be related to the dominance of different forces and interactions in the sludge flocs, e.g., EPS enmeshment in the 33d-SRT sludge versus cationic bridging in the Mgenriched sludge. A dense EPS layer covering much of the surface of flocs and less hydrated flocs have been observed in sludge at high SRTs of 16 and 20 days [52]. The tighter structure from the Mg-enriched flocs and the flocs at an SRT of approximately 33 days could have prevented the penetration of urea and EDTA, acting as a physical barrier against floc disruption before and after the shift due to their chaotropic and chelating effects, respectively. This would explain why the sludge dissociation constants from the Mg-SBR and SRTSBR were not significantly affected by the washings with urea and EDTA solutions. Sludge flocs at SRTs of 16 and 20 days have been reported to be only slightly disturbed by EDTA and urea solutions due to a compact floc structure and a compact layer of EPS encapsulating the flocs, probably with high lipid content [31].
5. Conclusions Activated sludge floc structural stability and surface properties were studied under temperature shifts from 30 to 45 ◦ C. Floc structural stability was assessed by subjecting sludge flocs to different chemical agents in suspending solu-
tions, and sludge surface charge was measured before and after the temperature shift. The main conclusions drawn from this work are: (1) The temperature shift from 30 to 45 ◦ C deteriorates the flocculating physicochemical properties of the sludge, as assessed by the consistent increase in the negative surface charge of activated sludge flocs. (2) Magnesium sludge enrichment and a high sludge retention time (SRT) of 33 days subdued the increase in the negative sludge surface charge in comparison with the increase in the negative sludge surface charge in the nonmanipulated reactor and promoted a less negatively charged sludge. Magnesium ions appear to neutralize the negatively charged groups on the sludge flocs, which supports the idea that magnesium ions foster bioflocculation via polymeric bridging. (3) Magnesium enrichment of the sludge and a higher SRT of 33 days promoted structurally stronger flocs, as assessed by lower dissociation constants under different chemical agents. Nevertheless, the more compact and stronger floc structure obtained by magnesium enrichment and an SRT of 33 days were unsuccessful in significantly decreasing sludge deflocculation under the temperature shift from 30 to 45 ◦ C. Particularly, the 33dSRT sludge flocs became weaker than the Mg-enriched flocs after the temperature shift, suggesting a structural difference between the Mg-enriched and the 33d-SRT sludge. (4) Manganese and iron appear to be released by the sludge under the temperature shift; however, this behavior remains unexplained. Further research is required to clarify the role of metal species in deflocculation and the origin of the increase in negative sludge surface charge resulting from the temperature shift from 30 to 45 ◦ C.
Acknowledgments The authors acknowledge financial support from the members of the Consortium “Minimizing the Impact of Pulp and Paper Mill Discharges” at the Pulp & Paper Centre, University of Toronto: Aracruz Celulose S.A., Carter Holt Harvey Tasman, Domtar Inc., Eka Chemicals Inc., GeorgiaPacific Corporation, Irving Pulp and Paper Limited, Japan Carlit Co. Ltd., ERCO Worldwide (formerly Sterling Pulp Chemicals, Ltd.), and Tembec Inc. In addition, financial support from the Government of Ontario/Du Pont Graduate Scholarship in Science and Technology is gratefully acknowledged, as well as partial support from the Natural Sciences and Engineering Research Council (NSERC) of Canada. The authors thank Amy Lo and Doug Craig at Domtar Inc. for help in obtaining mill effluent samples.
F. Morgan-Sagastume, D.G. Allen / Journal of Colloid and Interface Science 281 (2005) 136–145
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