Ind. Eng. Chem. Res. 2000, 39, 2221-2227
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Sodium Dodecylbenzenesulfonate Removal from Water and Wastewater. 2. Kinetics of the Integrated Ozone-Activated Sludge System Fernando J. Beltra´ n,* Juan F. Garcı´a-Araya, and Pedro M. A Ä lvarez Departamento de Ingenierı´a Quı´mica y Energe´ tica, Facultad de Ciencias, Universidad de Extremadura, 06071 Badajoz, Spain
A batch-activated sludge process (AS) has been applied to eliminate sodium dodecylbenzenesulfonate (NaDBS), a linear alkylbenzenesulfonate (LAS) compound, from synthetic and real domestic wastewater. The kinetics of surfactant biodegradation was best described by firstorder kinetics with rate constants of 1.28 and 1.15 L‚(g VSS‚day)-1 for synthetic and real domestic wastewater, respectively. The effect of preozonation on the overall surfactant and COD removal rates was also studied. The results indicated that treatment by combined ozone-AS leads to negligible foam ability and low residual surfactant concentration and COD in the treated effluent. Acclimation of a mixed culture to ozonated products was beneficial to highly improve biodegradation rates after preozonation. Thus, when applying ozone at a dose of 100 mg‚L-1, the first-order surfactant biodegradation rate constant increased up to 1.79 and 3.09 L‚(g VSS‚day)-1 for synthetic and real wastewater, respectively. Continuous experiments of ozonation followed by activated sludge of synthetic wastewater were also carried out, and the reactor performances were compared to those obtained from the application of the available kinetic data (derived from batch experiments) to the operating conditions. Good agreement between experimental and calculated data confirmed the reliability of the model. Introduction World consumption of linear alkylbenzenesulfonates (LASs) in 1996 was estimated around 2.3 tons, being the second surfactant in importance behind soap.1 Because of this widespread use of LAS, special attention has been paid to the occurrence and fate of LAS in the environment.2,3 As the main route of surfactant disposal is by municipal wastewater treatment, the development of methods directed at LAS removal from domestic wastewater, avoiding its presence in effluents released to surface water, is of great interest. In the first part of this work the ozonation kinetics of a LAS model compound, sodium dodecylbenzenesulfonate (NaDBS), was studied. It was concluded that high ozone doses are required to completely remove the surfactant from domestic sewage due to other organics present that compete for the available ozone. Therefore, ozonation alone is not recommended as a specific treatment to remove LAS from wastewater. Nevertheless, among the advantages of ozonation in wastewater treatment, the ozone ability to improve biodegradability, due to the transformation of nonreadily biodegradable compounds to lower molecular weight substances more amenable to microorganisms, can be highlighted. This approach requires a biological stage following the ozonation.4 The ozone dose for optimum biodegradation of the transformed products has to be carefully determined because the process becomes economically unsuitable when too high oxidant doses are applied. A number of authors have reported that alkylbenzenesulfonates are not completely destroyed by ozonation but are converted to nonsurface active and readily biodegradable substances.5 Then, the goal of surfactant ozonation should be the removal of negative effects such as foaming and color and the increase of wastewater
biodegradability rather than the complete breakdown to CO2 and H2O that can be achieved following biological oxidation. The objectives of this second part of the work were (1) to establish the biodegradation of NaDBS in domestic sewage; (2) to study the effect of preozonation on the kinetics of aerobic biological oxidation of municipal wastewater containing NaDBS, through batch experiments; and (3) to test the validity of the ozonation and biological oxidation kinetic parameters in a continuous integrated ozone-activated sludge (AS) system. Experimental Section Materials. NaDBS obtained from Aldrich was mixed either with domestic sewage collected from the municipal wastewater treatment plant at Badajoz (Spain), after being subjected to primary clarification and filtrated in the laboratory thereafter, or with synthetic wastewater prepared in the laboratory as described in the first part of this work.6 The mixed activated sludge culture used for biodegradation was taken from the aeration basin of the Badajoz municipal wastewater treatment facility and was acclimated to NaDBS wastewater for a period of 2 weeks prior to experiments. Experimental Setup and Procedure. Ozonations were conducted in the experimental device already described in part 1.6 Batch biodegradation was performed with activated sludge in a 4 L glass standard agitated tank operating in fill and draw mode. Acclimated sludge was centrifuged and mixed with wastewater to produce a suspension of approximately 1.5 g‚L-1 of mixed volatile suspended solids (MLVSS) per liter. At the beginning of each experiment the reactor was fed with 2.5 L of this suspension. Air was supplied
10.1021/ie9907223 CCC: $19.00 © 2000 American Chemical Society Published on Web 06/13/2000
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Figure 1. Schematic flow diagram of the continuous ozone-activated sludge system: 1, reservoir for wastewater; 2, ozonation column; 3, aeration basin; 4, clarifier.
through two diffusers placed at the bottom of the reactor in order to keep the dissolved oxygen concentration (DO) constant at 3 mg‚L-1. The reactor operated at a constant temperature of 20 ( 0.5 °C, and mixing was provided by mechanical paddles. At determined time intervals, a 40 mL sample was withdrawn from the reactor, centrifuged, and analyzed for methylene blue active substances (MBASs) and COD. Also, temperature, DO, MLVSS, and pH were recorded. Most of the experiments were performed on ozonated wastewater. In these cases preozonation was carried out as described in part 1 of this work6 and the wastewater was then biologically treated as described above. For continuous integrated ozone-AS the experimental setup shown in Figure 1 was used. The ozonation reactor was first charged with 1.5 L of the feed wastewater. Then, oxidant gas and feed wastewater were continuously supplied at 30 and 3.6 L‚h-1, respectively. When the bioreactor became three-fourths full, 2.5 L of the acclimated sludge (VSS: 10 g‚L-1) was loaded. The hydraulic retention time within the bioreactor was set at 5 h, and the return activated sludge pump capacity was controlled at 100% of the wastewater flow rate for all runs. A mixed liquor of volatile suspended solids (MLVSS) was kept constant at approximately 1.5 g‚L-1 by removing periodically the excess of sludge from the aeration basin. The temperature and DO in the aeration basin were also kept constant at 3.0 mg‚L-1 and 20 °C, respectively. Attainment of steady-state was assumed when analysis of samples withdrawn at different times from the digester gave constant results in consecutive COD measurements. Analytical Methods. Standard methods were employed for the characterization of wastewater.7 The methylene blue active substance (MBAS) was used as an indicator of anionic detergent in wastewater (i.e. NaDBS) while COD was measured in the liquid samples following the dichromate reflux method. Ozone in water was analyzed by the Indigo method8 while ozone in the gas entering and leaving the reactor was measured with an Anseros Ozomat GM-109 analyzer. DO was recorded
on an oxygen meter (YSI 58), and MLVSS was followed with a Dr. Lange turbidimeter provided with a HT1 probe. Results and Discussion Biodegradation Kinetics. Several batch experiments were first conducted in synthetic and real domestic wastewater to investigate the kinetics of biodegradation by AS. NaDBS was added to both types of wastewater so that the MBAS concentration was about 15 mg‚L-1. It must be noted that the real wastewater had a MBAS background (i.e. MBAS concentration before the addition of NaDBS) between 3 and 5 mg‚L-1, which represented 20-30% of the total MBAS concentration once NaDBS was added. Regardless of the nature of the wastewater, it was found that >60% of the surfactant and COD were removed after 24 h of treatment with an average MLVSS concentration of 1.5 g‚L-1. However, residual surfactant and COD remained after a long time of biodegradation. On the basis of previous works, first-order degradation kinetics for both anionic surfactant and COD was assumed to determine the biodegradation rate constants.9-11 At constant biomass and limiting substrate levels, the rate of substrate consumption therefore can be assumed to be directly proportional to the substrate concentration, S:
dS ) kX(S - S*) dt
(1)
where S* is the concentration of substrate not biodegraded after a long time of bioassay. After rearranging, the integrated form of eq 1, considering X as constant, is
S ) S* + (S0 - S) exp(-k′t)
(2)
where k′ ) kX. Figure 2 shows the plot of concentration of MBAS and COD against the reaction time, t. The
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Figure 2. Removal of MBAS and COD during domestic wastewater biodegradation: a, synthetic wastewater; b, real wastewater; 0, MBAS; O, COD.
results presented correspond to biological oxidation of synthetic and domestic wastewater where MBAS and COD have been considered as substrate for the kinetic study. The experimental data are fitted to eq 2 (see solid lines), and good agreement is observed between the experimental and calculated data, which confirms the kinetics assumed. The concentration of microorganisms, measured as volatile suspended solids (X), could be considered constant, since it did not vary more than 10% throughout the experiments. Thus, from the nonlinear regression analysis of experimental data, k and S* were calculated for MBAS and COD. For the surfactant biodegradation rate constant, kMBAS, values of 1.28 and 1.16 L‚(g VSS‚day)-1 were found for synthetic and real wastewater, respectively, while kCOD was calculated to be 5.59 and 5.65 L‚(g VSS‚day)-1 for synthetic and real wastewater, respectively. On the other hand, the nonbiodegradable fraction of each substrate may be defined as the ratio S*/S0. On the basis of the former assumption
and by considering the experimental results presented in Figure 2, about 19% and 27% of surfactant and COD, respectively, cannot be removed by AS from synthetic wastewater while approximately 18% and 35% are the percentages of surfactant and COD that remained after AS treatment of real domestic wastewater. It is worthy of note that the amount of residual MBAS concentration was higher than that from the literature where the AS system was applied for domestic wastewater treatment.9-11 This is likely due to the high initial MBAS concentration used in this work compared with other studies. The similar C*MBAS and kMBAS values for synthetic and real domestic wastewater suggest that the surfactant biodegradation does not depend on the type of wastewater while differences in COD* for such kinds of wastewater can be due to differences in the concentration of readily biodegradable compounds within the real and synthetic wastewater. Effects of Preozonation. (i) Foaming Ability. During individual chemical and biological treatment of NaDBS, the foaming ability of the wastewater was observed. Thus, at the beginning of wastewater ozonation containing 15 mg‚L-1 of NaDBS, a layer of foam was produced, but as the reaction proceeded, the foaming ability was drastically reduced. With regard to AS treatment, foaming was observed in the aeration basin when treating nonozonated wastewater. On the contrary, the foaming ability of the wastewater disappeared when aeration at the biological stage was applied just after ozonation. Ozonation favored oxygen transfer during the AS process and allowed a high microorganism respiration rate to be achieved and allowed us to maintain a dissolved oxygen (DO) concentration of about 3 mg‚L-1 with an air flow rate lower than that needed in a conventional treatment without preozonation. (ii) Acclimation Stage. As a result of wastewater ozonation, new products are expected to be formed. Very often the mixed culture in activated sludge requires time to readapt to some of the ozonated products. In these cases, an acclimation period is required.12 In this work, repeated experiments of biological oxidation of preozonated domestic and synthetic wastewater were carried out. Experiments were performed daily, operating in fill and draw mode for a whole period of 1 month in order to examine the ability of different-time-acclimated microorganisms to remove anionic surfactant and COD from preozonated wastewater. Figure 3 shows the variation of surfactant concentration and COD with reaction time for some experiments carried out with synthetic wastewater. It is deduced that surfactant and COD removal rates during the first week showed biodegradation rates slower than those achieved after >1 week of acclimation. Longer periods of acclimation did not improve biodegradation rates and overall COD removals. Following the nonlinear regression procedure described before, the biodegradation rate constants, kMBAS and kCOD, were determined. When treating preozonated wastewater with nonacclimated bacteria, rate constants were found to be even lower than those achieved without preozonation. However, when using >1 week-acclimated microorganisms, rate constants reached maximum values. (iii) Ozone Dose. The effect of preozonation at different ozone doses on the subsequent biological oxidation to reduce the concentration of surfactant and the COD levels of domestic wastewater was also studied through first-order kinetic parameters of biodegrada-
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Table 1. First-Order Kinetic Parameters of Synthetic and Real Domestic Wastewater Biodegradation: Influence of a Preozonation Stage type of wastewater synthetic
real
ozone dose, mg‚L-1
kMBAS, L‚(g VSS‚day)-1
C*MBAS/CMBAS0, %
kCOD, L‚(g VSS‚day)-1
COD*/COD, %
a 50 100 150 200 a 50 100 150 200
1.28 1.60 1.79 1.89 1.72 1.15 1.95 3.09 3.41 2.72
19.0 9.4 6.0 1.6 0.5 18.3 17.9 13.7 13.4 11.3
5.59 6.17 6.49 6.20 6.83 5.65 6.66 5.70 5.97 6.05
26.4 26.5 19.4 16.7 16.5 35.0 29.4 25.0 26.5 24.2
a Nonozonated wastewater. Biodegradation conditions: T ) 20 °C; pH ) 7-8; MLVSS ) 1.5-1.6 g‚L-1; C -1 MBAS ) 12-15 mg‚L ; COD0 ) 275-310 mg‚L-1.
Figure 3. Influence of microorganism acclimation time on surfactant and COD removal by activated sludge: a, variation of MBAS as mg‚L-1 of NaDBS; b, variation of COD with time. Conditions: T ) 20 °C; pH ) 7; reaction volume ) 2.5 L; DO ) 3 mg‚L-1; MLVSS ) 1.5 g‚L-1; ozone dose applied at the pretreatment stage ) 100 mg‚L-1. Acclimation time: 0, nonacclimated; O, 1 week; 4, 2 weeks; 3, 3 weeks.
tion. Thus, from Table 1 it is seen that the preozonation stage at an ozone dose between 50 and 200 mg‚L-1
enhanced the surfactant and COD biodegradation rates with respect to those of the process treating nonozonated wastewater. Also, the concentration of remaining MBAS and COD fractions after biodegradation (i.e. C*MBAS and COD*) markedly diminished with increasing ozone dose up to 200 mg‚L-1, especially in the case of synthetic wastewater. This improvement in biodegradation rates and residual biorefractory compounds after biotreatment can be explained because of the nature of the ozonation products, probably ozonides, formaldehyde, and organic acids.5 These products seem to be amenable to biological break down, and therefore they are easily removed at the subsequent biological oxidation stage. Nevertheless, further data about the identification of intermediate and final ozonation products and their biodegradability are needed to confirm this explanation. Integrated Continuous Ozone-AS Process. Kinetic models are required for the design of the combined ozonation-AS process and the determination of the optimum operating conditions in terms of efficiency and economic cost. In this sense, values of reaction rate constants of NaDBS derived from batch experiments were used to predict the reactor performances under continuous conditions and to check the proposed kinetic model. In addition, the nonideal behavior of wastewater and gas flowing through the continuous reaction system was studied and it was later considered in the kinetic model. (i) Deviation from Ideal Flow Conditions of the Actual Continuous Reaction System. Fluid dynamics of phases flowing through reactors can strongly influence the mass-transfer and chemical reaction rates and, therefore, the kinetics of such a process. Thus, residence time distribution (RTD) experiments were first carried out in order to asses the flow behavior through both actual ozonation and AS reactors13 (see Figure 1 for experimental setup details). For so doing, ozone and potassium hydrogen phthalate were used as tracers in the carrier gas and liquid, respectively. These experiments showed some deviation from the ideal flow conditions (i.e., perfectly mixed or plug flow behavior). Therefore, experimental models for nonideal flow situations were applied. Thus, residence time distribution (RTD) curves were analyzed by the completely mixed reactor in series model.13 For the liquid-phase behavior study, the wastewater effluent response from the ozonation reactor to step inputs led us to simulate the behavior of the actual reactor by two identical stirrer tank reactors in series whose total liquid volume is the same as that of the actual reactor. On the other hand, the gas flow presented the same residence time distribution function that would occur within eight stirrer
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Figure 4. Schematic diagram of the proposed continuous flow model for the combined ozonation-AS experimental system.
tank reactors in series. Eventually, the evaluation of the flow conditions in the activated sludge tank led to a stirrer tank reactor with an effective volume, VAS, of 17.2 L. The remaining volume to complete the total reaction volume (i.e., 0.8 L) was considered to be a stagnant region. Accordingly, Figure 4 shows the scheme for the proposed flow model for the combined ozonationAS system through a series of ideal perfectly mixed reactors that simulates the flow behavior of wastewater and gas flowing through the actual reaction system. It should be noted that gas and liquid flowed countercurrently during ozonation. Therefore, according to the model proposed in Figure 4, wastewater enters reactors 1-4 while the gas enters reactor 8. (ii) Reaction Model. Degradation of NaDBS by the combined ozone-AS process can be summed up as a series of reactions.
-rNaDBS ) -
dCNaDBS kd ) CO3CNaDBS + dt z kHO•CHO•CNaDBS (8)
Mole balance of ozone in water (reactors 1-8) kLa
[
]
CO3g F F - CO3 βV + CO30 - CO3 + rO3V ) 0 He 4 4
(9)
NaDBS + zO3 f products
(3)
O3 + OH- f HO• + other free radicals
(4)
In eq 9 kLa is the liquid-phase volumetric masstransfer coefficient, which was found to be 1.76 × 10-2 s-1 by means of a series of oxygen absorption experiments in synthetic wastewater;14 He, the Henry’s law constant, determined following a procedure already described,15 resulted in a value of 75 atm‚L‚mol-1 at 20 °C and pH 7.5; the liquid hold-up within the bubble ozonation column, β, was found to be 0.95 at the steadystate; finally, rO3 is the decomposition rate term of ozone due to chemical reactions, not only with NaDBS but also with any other dissolved compound. A pseudo-first-order kinetics was therefore considered for ozone decomposition:16
NaDBS + HO• f products
(5)
-rO3 ) kO3CO3
Ozonation:
Biodegradation: NaDBs + X + O2 f products
(6)
Modeling of O3/AS oxidation of NaDBS in wastewater can be accomplished by solving the set of mole balance equations of species present (for notation see Figure 1 and Nomenclature Section). For each of the assumed well-mixed reactors of Figure 4, the mole balance equations are as follows for the continuous steady-state ozonation:
Mole balance of NaDBS (reactors 1-8) F F - C + rNaDBSV ) 0 C 4 NaDBS0 4 NaDBS(i)
(7)
where the rate of NaBDS decomposition by ozonation is mathematically described as follows
(10)
Mole balance of ozone in the gas (reactors 1-8) FgCO3g0 - FgCO3g + kLa
[
]
CO3g - CO3 βV ) 0 (11) He
For the continuous steady-state AS process:
Mole balance of NaDBS (AS reactor) FCNaDBS0 - FCNaDBSf + kNaDBSCNaDBSfXVAS ) 0 (12) where first-order kinetics for the specific biodegradation of NaDBS was considered. It may be noted that the first-order biodegradation rate constant, kMBAS, calculated in this work and the concentration of MBAS, CMBAS, have been used as kNaDBS and CNaDBS throughout
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Figure 5. Flow chart of the algorithm used for solving the combined ozonation-AS kinetic model. Table 2. Calculated and Experimental Values of NaDBS and Ozone Concentrations for the Continuous Combined Ozonation-AS Processa experimental data
calculated data
CO3g0
CO3
CNaDBS
CNaDBSf
CO3g0
CO3
CNaDBS
CNaDBSf
kO3
run no.
mg‚L-1
mg‚L-1
mg‚L-1
mg‚L-1
mg‚L-1
mg‚L-1
mg‚L-1
mg‚L-1
h-1
1 2
5.1 7.3
0.6 0.78
12.3 10.4
9.2 5.2
5.1 7.3
0.48 0.66
12.5 10.8
9.7 6.0
174 181
a Experimental conditions. Ozonation run no. 1: T ) 20 °C; pH ) 7; F ) 30 L‚h-1; F ) 3.6 L‚h-1; C -1 g O3g0 ) 5.1 mg‚L ; CO3g ) 2.1 mg‚L-1; V ) 1.5 L; CNaDBS0 ) 15.0 mg‚L-1. Run no. 2: T ) 20 °C; pH ) 7; Fg ) 30 L‚h-1; F ) 2.0 L‚h-1; CO3g0 ) 7.3 mg‚L-1; CO3g ) 2.9 mg‚L-1; V ) 1.5 L; CNaDBS0 ) 15.0 mg‚L-1. AS run no. 1: T ) 20 °C; pH ) 6.8-7.3; MLVSS ) 1.5-1.6 g‚L-1; DO ) 3-4 mg‚L-1; VAS ) 18 L. Run no. 2: T ) 20 °C; pH ) 7.0-7.2; MLVSS ) 1.4-1.6 g‚L-1; DO ) 3-4 mg‚L-1; VAS ) 18 L.
eqs 7-12 because NaDBS was the main surfactant in the wastewater employed. (iii) Solution of the Kinetic Model. To validate the kinetic model of the reaction system, the set of linear equations from the mole balances within each reactor of Figure 4 must be solved and the values of the concentrations of the final effluents must coincide with those from the experimental results. Figure 5 shows the algorithm of the mathematical procedure followed to solve the system of equations (eqs 7-12) applied to each reactor. The ozonation stage was first studied. Experimental fluid dynamics, concentration of species (i.e. CNaDBS0, CO3g, and CHO•), and rate constant data (the latter derived from kinetic study of batch experiments6) were provided to start the calculation program. To estimate the concentration of hydroxyl radical (an unknown variable of the system), semibatch experiments were conducted at constant ozone mass flow rate and initial MBAS concentration, to reach a MBAS conversion similar to that achieved by the continuous
running ozonation. Thus, HO• was estimated from these experiments as follows:
CHO• )
∆CNaDBS kd - CO3CNaDBS ∆t z kHO•CNaDBS
(13)
Following this procedure, CHO• was found to be 1.1 × 10-14 M for the conditions of experiments 1 and 2 of Table 2. Following the calculation scheme showed in Figure 5, a value of the ozone decomposition rate constant, kO3, was also assumed and later checked using a trial and error method. As a result of the overall procedure, values of CNaDBS, CNaDBSf, and CO3 were obtained and could be compared with experimental results. As shown in Table 2, deviations between experimental and calculated effluent concentrations were
