Water Science & Technology sample issue by IWA Publishing

70.1

2014

Thermal hydrolysis of waste activated sludge at Hengelo Wastewater Treatment Plant, The Netherlands Mathijs Oosterhuis, Davy Ringoot, Alexander Hendriks and Paul Roeleveld

ABSTRACT The thermal hydrolysis process (THP) is a sludge treatment technique which affects anaerobic biodegradability, viscosity and dewaterability of waste activated sludge (WAS). In 2011 a THP-pilot plant was operated, connected to laboratory-scale digesters, at the water board Regge en Dinkel and in cooperation with Cambi A.S. and MWH Global. Thermal hydrolysis of WAS resulted in a 62% greater volatile solids (VS) reduction compared to non-hydrolysed sludge. Furthermore, the pilot digesters could be operated at a 2.3 times higher solids loading rate compared to conventional sludge digesters. By application of thermal sludge hydrolysis, the overall efﬁciency of the sludge treatment process can be improved. Key words

| sludge digestion, thermal hydrolysis, waste activated sludge

Mathijs Oosterhuis (corresponding author) Water Board Regge en Dinkel, currently: Water Board Vechtstromen, P.O. Box 5006, 7600 GA, Almelo, The Netherlands E-mail: m.oosterhuis@vechtstromen.nl Davy Ringoot Cambi A.S. P.O. Box 78, NO-1371 Asker, Norway Alexander Hendriks Royal HaskoningDHV, currently: Delft University of Technology, P.O. Box 5048, 2600 GA, Delft, The Netherlands Paul Roeleveld MWH Global, currently: RoyalHaskoningDHV, P.O. Box 151, 6500 AD, Nijmegen, The Netherlands

INTRODUCTION Wastewater sludge management in the Netherlands is controlled by strict regulations for land application. Sludge digestion is a widespread technology which is used for the reduction of sewage sludge and the production and recovery of energy as biogas. Biogas is usually converted into electricity which, in most cases, is used at the wastewater treatment plant (WWTP) itself. Primary sludge (PS) is easily biodegradable and results in a volatile solids (VS) removal of 50 to 60% and high biogas yields. Waste activated sludge (WAS) from WWTPs with low sludge loading rates ( 0.05 kg biochemical oxygen demand (BOD)/kg volatile suspended solids (VSS).d) and a long solids retention time (SRT), has lower biodegradability in sludge digesters and typical VS removal rates are 25 to 50% (Bolzonella et al. ; STOWA b). The focus for the use of the thermal hydrolysis process (THP) is on increasing the energy recovery from sludge. The possible improvement in sludge dewatering by use of THP is an adjunct advantage. Experiments with sludge disintegration (ultrasonic) at several sludge digesters in the Netherlands did not result in an expected increase of VS removal, despite promising results in Germany (Luning et al. ; STOWA ). This experience provided the insight that results with WAS from one country cannot be converted to the same expectations in another country. This depends, for example, on a difference doi: 10.2166/wst.2014.107

in wastewater quality, design standards and efﬂuent requirements. Therefore, verifying the potential of sludge reduction technologies for the particular type of sludge is recommended. Heating of WAS up to temperatures of 150–190 C results in solubilisation of chemical oxygen demand (COD) and consequently higher VS removal rates when the sludge is digested (Li & Noike ; Bougrier et al. ). Thermal hydrolysis also effects the viscosity of sludge. It decreases substantially, which makes it possible to feed the digesters with a higher VS concentration. The ﬁrst full-scale results with thermal sludge hydrolysis date from 1995 in Norway, Denmark and in the UK (STOWA a). An additional advantage of the process is that the sludge is sterilised, which makes land application possible as fertiliser in the form of biosolids (class A) in several countries. The Dutch legislation for land application is, however, too strict to allow the application of sterilised sludge as a fertiliser in agriculture. Although land application of sewage sludge is not allowed in the Netherlands, the extra energy production, increased anaerobic reduction of VS and the expected improvement of sludge dewatering mean that application of thermal hydrolysis of WAS is still an attractive option. The water board Regge en Dinkel is considering centralising the sludge treatment of 800,000 population W

M. Oosterhuis et al.

Thermal hydrolysis of waste activated sludge at WWTP

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2014

equivalents at one location (Hengelo) in combination with thermal sludge hydrolysis. Expensive revisions of old sludge digesters at Enschede could be avoided in this way and the efﬁciency of sludge treatment could be further improved. The feasibility of thermal sludge hydrolysis was evaluated in a pilot plant study in 2011. The aim of the study was to determine the effect of THP on the anaerobic biodegradation and dewaterability of WAS.

METHODS Dewatered WAS (16% total solids (TS)) was treated in a Cambi pilot plant. The sludge was heated by steam injection for 20 min until it reached 165 C and 6 bar. After 20 min the pressure was lowered to atmospheric pressure by opening a valve, resulting in steam explosion of the sludge. Every week, batches of 20 kg WAS were hydrolysed (Figure 1). To simulate the full-scale operation mode of the digesters, hydrolysed sludge was mixed with PS in a ratio of 80% WAS and 20% PS (based on total solids content). The mixtures were stored for a maximum of 1 week at 4 C in 450 mL bottles. The sludge mixtures were digested in four small digesters (8 litre volume each) (Figure 2). Digesters 1 and 2 were fed with a mixture of hydrolysed WAS and PS, and digesters 3 and 4 were fed with a mixture of untreated WAS and PS (80/20%). The feed for digesters 3 and 4 was also prepared weekly. Every day, the digesters were sampled and fed manually, always in the same order between 8.00 and 10.00 a.m. To avoid temperature shocks, the sludge was heated in a microwave to 30 C before being added to the digesters. W

Figure 2

Pilot sludge digesters at Hengelo WWTP.

All digesters were operated at a temperature of 37 C and a sludge retention time of 20 days. The digesters that were fed with hydrolysed sludge (1 and 2), were started up with digested sludge from a full-scale sludge digester in Denmark where WAS is treated with the THP process. The average sludge loading rate of digesters 1 and 2 was 3.9 kg VS/(m³ digester.day) and the average sludge loading rate of digesters 3 and 4 was 1.7 kg VS/(m³ digester.day). The thermal hydrolysis pilot plant and digesters were tested during the 6 months from May to November 2011. Biogas production was measured with gas clocks, and twice a week, digested sludge was sampled and analysed for total solids, ash content, COD, P-tot, NH4-N and pH, according to standardised analysis methods: NEN-EN 12880 (TS), NEN-EN 12879 (ash content), NEN 6633 (COD), NEN 6645, NEN-EN 15681-2 (total P), NEN-646 (NH4-N). Table 1 summarises the experimental setup.

Table 1

Experimental setup of the digesters

Volume (litre)

Figure 1

Pilot thermal sludge hydrolysis (Cambi) at Hengelo WWTP.

Digesters 1 & 2

Digesters 3 & 4

Average feed sludge mixture

10.4% TS

4.5% TS

WAS

Hydrolysed

Non-hydrolysed

Sludge loading rate (kg VS/(m³.day))

3.9

1.7

Sludge retention time

20 days

Temperature

37 C

20 days W

37 C

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Thermal hydrolysis of waste activated sludge at WWTP

As well as the sludge pre-treatment and operation of the four digesters, the following additional laboratory-scale experiments were carried out:

• •

The foaming potential and stability of hydrolysed sludge was tested at laboratory scale. The test protocol was based on a description in Hug (). Sludge dewatering was tested by thermogravimetric measurements at laboratory scale (Kopp & Dichtl ).

RESULTS Figure 3 presents the monthly average speciﬁc biogas productions in litre biogas/kg VS. For this, gas production for the individual digesters was averaged. The data make clear that the difference in speciﬁc biogas production between

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digesters fed with hydrolysed sludge and non-hydrolysed sludge was 30–40%. The variation in average biogas production between the individual digesters was relatively low at 3.3% (THP) and 2.6% (reference). Furthermore, speciﬁc biogas production decreased during the test in all digesters. This can be explained by decreasing anaerobic biodegradability of the WAS during summertime due to higher aerobic mineralisation of the WAS. Because of long SRTs (>20 days) in the digesters, in combination with the storage of sludge, this decrease in VS removal is visible until October. The decrease in speciﬁc biogas production is even greater in the reference digesters. This suggests that thermal hydrolysis is more effective in summertime than in wintertime. This is in line with Bougrier et al. () who found a similar effect from THP. Since mixtures of PS and WAS were digested to simulate the planned operation mode of the digesters in full scale, the absolute effect of THP on the VS removal of WAS could not be determined exactly. Therefore, we assumed an average VS removal of 55% for PS and estimated the increase in VS removal of WAS due to thermal hydrolysis. The results of this calculation are presented in Table 2. Table 2 shows that the estimated VS removal of WAS increased from 26 to 42%, which is an increase of 62%. Pilot experiments with the THP-process of Sustec at the Venlo WWTP in the Netherlands showed a 55% increase of VS removal after thermal hydrolysis of WAS at a temperature of 140 C (van Dijk & de Man ). Pilot experiments at the Amersfoort WWTP in the Netherlands with Sustec (STOWA ) showed an increase in VS removal of WAS of 38% at 140 C. The WAS in our study had a relatively low VS content (67–70.0%). This indicates a high degree W

Figure 3

Table 2

Monthly average biogas production (l/kg VS feeding to the digester).

Estimation of VS removal of WAS (14-6-2011–8-11-2011)

Out

Removal

[g VS/d]

[%]

PS (assumed) Reference

2.9

1.31

1.60

55%

THP

6.6

2.97

3.63

55%

Reference

10.9

8.10

2.81

26%

THP

24.5

14.13

10.37

42%

WAS sludge (estimated)

Total (measured) Reference

13.8

9.4

4.4

31.9%

THP

31.1

17.1

14.0

45%

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Thermal hydrolysis of waste activated sludge at WWTP

of aerobic mineralisation of the sludge, which could explain the high effect of THP (Bougrier et al. ). Performance of the digesters The pH in the reference digesters was about 7.5 while the pH in the THP digesters was about 8. The higher pH could be explained by high ammonium concentrations in the THP digesters. A high loading rate for the THP digesters and increased VS removal resulted in ammonium concentrations of 3–4 g NH4-N/L while the ammonium concentration in the reference digesters was about 1 g NH4-N/L. Inhibition of methanogenesis due to toxic NH3 would have resulted in increased levels of volatile fatty acids (VFA) but this was not observed. VFA concentrations in the THP digesters were much higher than in the reference digester, but stable. See supplementary information in Table S1 for average concentrations (available online at http://www.iwaponline. com/wst/070/107.pdf). That no inhibition of the digestion process occurred in the THP digesters is remarkable, since other studies report inhibition at THP temperatures of 175 C and higher (Haug et al. ). In the study of Haug et al. (), dilution of the sludge until 4% TS was necessary to avoid inhibition of the digestion process. In our pilot plant, average total solids concentrations in the sludge feed were 10.4% TS (see Table 1). That no inhibition of the digestion occurred is probably caused by the use of inoculum anaerobic sludge from a full-scale digester in Denmark, where WAS is treated with the THP process. Adaptation to inhibiting compounds in THP sludge is necessary when starting up digesters with THP sludge (Stuckey & McCarty ). W

Foaming Foaming was observed in all digesters, however, the digesters with hydrolysed sludge seemed to foam more than the reference digesters. This effect could be explained by higher VFA concentration and biogas production. Both parameters are related to foaming problems in digesters (STOWA ). A sludge foaming test was executed to determine foaming potential and foam stability. The test protocol was based on a description in Hug (). At normal gas velocities, foaming potential and foam stability are not signiﬁcantly higher when hydrolysed sludge is digested (see supplementary information Table S2 for the test results; available online at http://www. iwaponline.com/wst/070/107.pdf). The gas velocity in a sludge digester is determined by the biogas production/m2 of the digester surface and the mixing power when mixing by biogas injection is applied. The biogas production will

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increase substantially when the THP process is applied. This might introduce foaming problems in digesters. However, due to a strong homogenisation effect of the thermal hydrolysis it is expected that less mixing power is required. The gas velocity in the digester does not, therefore, necessarily increase after implementation of the THP process. Biogas composition Biogas was sampled and analysed twice, but no signiﬁcant differences in CH4 concentrations were found. The average CH4 concentration in biogas was about 65% in the THP digesters and the reference digesters. Sludge dewatering test The sludge dewaterability was tested with a thermogravimetric measurement according to Kopp & Dichtl (). The results of the dewatering test are presented in Table 3. The dewatering tests make clear that better sludge dewatering results can be achieved when thermal sludge hydrolysis is applied. However, the polymer demand for sludge dewatering seems to increase substantially. It should be mentioned that another dewatering test by a polymer supplier showed a speciﬁc polymer demand of 13 kg poly electrolyte (PE)/ton TS which is much lower than the 18 kg PE/ton TS that was found in this dewatering test. Addition of aluminium to the sludge with a Metal:Phosphorus ratio of 1 (mol/mol) resulted in a decreased polymer demand. In general, the type of polymer and addition of ferric or aluminium chloride will affect the total polymer demand and dewatering result. More data are required to determine more accurately the polymer demand for dewatering of THP sludge. Although the sludge dewatering was tested at laboratory scale, the test result compared well to the full-scale sludge Table 3

Results of thermogravimetric dewatering tests

% TS

Polymer demand (kg active polymer/ton TS)

Reference sludge Hengelo WWTP

24.1a

10.0

Reference digested sludge pilot

26.0

10.5

Hydrolysed sludge pilot

33.9

18.0

Hydrolysed sludge pilot þ aluminium addition

34.5

16.0

This result corresponds with the full-scale dewatering at the Hengelo WWTP: 23.3% d.s.

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Thermal hydrolysis of waste activated sludge at WWTP

dewatering. The reference sludge of the full-scale dewatering centrifuge at the Hengelo WWTP had a TS content of 23.3%, which is 0.8% lower than the prediction for the test (24.1%). The thermogravimetric dewatering test seems to be quite accurate. However, only one sludge sample from the pilot digesters was tested after 4 months of operation. These results should therefore be interpreted carefully; longer testing of sludge dewatering is necessary to obtain conclusive results. Other pilot research in the Netherlands showed similar sludge dewatering results. Pilot tests by Sustec on the Venlo and Amersfoort WWTPs showed that sludge could be dewatered until 31% and 25% TS, respectively, when thermal pretreatment was applied (van Dijk & de Man ; STOWA ).

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However the orthophosphate concentrations in reject water from the reference digesters would have been about 100–150 mg P/L, which are common concentrations in reject water after conventional digesters for WAS, from a WWTP with enhanced biological phosphorus removal. Compared to this, the concentrations when using THP are relatively low, especially taking into account the high sludge loading rate and the increased VS removal. The relatively low orthophosphate concentrations in the reject water could be explained by phosphate precipitation in the sludge; at a pH of 8, phosphates can precipitate as struvite (Jeong ).

GENERAL DISCUSSION Composition of reject water Thermal pre-treatment of WAS can result in an increase of soluble inert COD (STOWA a, ). The amount of inert COD was estimated in digested sludge from the pilot at the Hengelo WWTP. A sludge sample from the digesters with pre-treated sludge was prepared and dewatered by the addition of poly-electrolyte. Soluble inert COD was deﬁned as the difference between COD and (BOD5/0.65) in the water phase, the reject water. The concentration of inert COD was 1,078 mg/L, which is equivalent to 11 kg COD/ton d.s. dewatered sludge. Application of the THP process on the Hengelo WWTP and digestion of 16,000 ton d.s. sewage sludge/y will result in an increase of 7 mg/L COD in the WWTP efﬂuent. It should be emphasised that this is a worst case scenario. Reject water, containing inert COD, is recycled to activated sludge tanks and part of the COD can still be removed by adsorption to the activated sludge or by biodegradation. Orthophosphate concentrations in reject water from the pre-treated sludge were 129–172 mg P/L. The orthophosphate concentrations in the reject water from the reference sludge were not measured, because only the orthophosphate concentrations in the pre-treated digested sludge are relevant for calculation of the potential additional orthophosphate removal costs in the future.

Table 4

The results of the pilot test in our study make it clear that thermal hydrolysis of Dutch WAS is significantly improving the VS removal in mesophilic sludge digesters. No inhibition of the digesting process was observed at TS concentrations of 10.4% TS in the feed. Furthermore, a higher sludge loading rate for the sludge digesters is possible, and sludge dewatering will certainly be improved. The results are in line with other studies (see Table 4). It should be mentioned that the required energy for the THP process (steam production) can have a negative impact on the total energy balance of sludge treatment. However, when a mixture of 40% PS and 60% WAS is digested, the amount of heat that will be generated by the co-generation of heat and electricity will be sufficient to produce steam for thermal hydrolysis of WAS alone. Although centralised sludge treatment can be beneficial, it should be emphasised that the sludge treatment becomes more complex. Therefore, the following important aspects should be considered carefully before implementation:

• • • •

Reject water treatment for nitrogen removal Type of energy recovery from biogas ○ co-generation, production of transport fuel or heat generation Sludge dewatering and selection of poly-electrolytes Logistics planning for sludge transport.

Effect of thermal pre-treatment of WAS

Treatment WAS sludge W

120–190 C W

60–100 C

Effect mesophilic sludge digestion

References

40–60% increase VS removal

Haug et al. (); Bougrier et al. (, ); Climent et al. ()

15–48% increase VS removal

Haug et al. (); Hiraoka et al. (); Barjenbruch & Kopplow ()

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Thermal hydrolysis of waste activated sludge at WWTP

The preferred technique for reject water treatment is side-stream nitrogen removal technology. The exact effects of high temperatures during the pre-treatment of sludge on side-stream nitrogen removal processes are still unknown. However, it is known that high temperature (>150 C) pretreated sludge can be toxic for the anaerobic digestion culture (Stuckey & McCarthy ). The potential toxicity for nitrogen side-stream removal should be considered carefully before implementation of high temperature pre-treatment of sludge (Tattersall et al. ). Further research should investigate the possible inhibiting or toxic effects of the reject water from pre-treated sludge on nitrogen side-stream removal. Energy recovery from biogas can be realised with different techniques. Depending on the local situation, one technique should be preferred above the others. Sludge dewatering will be improved, but the type of polymer will have a great impact on ﬁnal dewatering results. Thermal hydrolysis of WAS is a promising technology that can make domestic wastewater treatment more energy efﬁcient. A few water boards in the Netherlands have already decided to implement this technology, and some water boards are considering implementation. Although the increase in VS removal and sludge dewatering is substantial, the economic feasibility also depends on local aspects. For Regge en Dinkel water board, the main trigger is to omit the expensive revision of an old sludge digester, which can be prevented when the THP process is implemented. The surplus of produced electricity should be delivered for a reasonable price to make centralised sludge treatment economically feasible. W

CONCLUSIONS The following conclusions can be drawn from the pilot research:

• • •

VS removal of WAS increased from 26 to 42% due to thermal hydrolysis, which is an increase of 62%. By application of the THP process, the sludge loading rate for sludge digesters could be 2.3 times higher compared to conventional sludge digesters. Thermal hydrolysed sludge can be dewatered to 30% TS.

ACKNOWLEDGEMENT The pilot test was partly ﬁnanced by STOWA.

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REFERENCES Barjenbruch, M. & Kopplow, O.  Enzymatic, mechanical and thermal pretreatment of surplus sludge. Adv. Environ. Res. 7 (3), 715–720. Bolzonella, D., Pavan, P., Battistoni, P. & Cecchi, F.  Mesophilic anaerobic digestion of waste activated sludge: influence of the solid retention time in the wastewater treatment process. Process Biochem. 40, 1453–1460. Bougrier, C., Delgenès, J. P. & Carrère, H.  Effects of thermal treatments on five different waste activated sludge samples solubilisation, physical properties and anaerobic digestion. Chem. Eng. J. 139, 236–244. Bougrier, C., Delgenès, J. P. & Carrère, H.  Combination of thermal treatments and anaerobic digestion to reduce sewage sludge quantity and improve biogas yield. Process Saf. Environ. Protect. 84 (B4), 280–284. Bougrier, C., Delgenès, J. P. & Carrère, H.  Impacts of thermal pre-treatments on the semi-continuous anaerobic digestion of waste activated sludge. Biochem. Eng. J. 34, 20–27. Climent, M., Ferrer, I., Baeza, M. D., Artola, A., Vazquez, F. & Font, X.  Effects of thermal and mechanical pretreatments of secondary sludge on biogas production under thermophilic conditions. Chem. Eng. J. 133, 335–342. Haug, R. T., Stuckey, D. C., Gossett, J. M. & McCarty, P. L.  Effect of thermal pretreatment on digestibility and dewaterability of organic sludges. J. Water Pollut. Control Fed. 73–85. Hiraoka, M., Takeda, N., Sakai, S. & Yasuda, A.  Highly efficient anaerobic digestion with thermal pretreatment. Water Sci. Technol. 17 (4–5), 529–539. Hug, T.  Characterization and controlling of foam and scum in activated sludge systems. PhD Thesis, Swiss Federal Institute of Technology, Zurich. Jeong, H.  Prediction of struvite formation potential in EBPR digested sludge using mass balances, batch precipitation tests, and chemical equilibrium modeling. Dissertation, University of Nevada, Las Vegas. Kopp, J. & Dichtl, N.  Prediction of full-scale dewatering of sewage sludge by thermogravimetric measurement of the water distribution. Water Sci. Technol. 42 (9), 141–149. Li, Y. Y. & Noike, T.  Upgrading of anaerobic digestion of waste activated sludge by thermal pretreatment. Water Sci. Technol. 26 (3–4), 857–866. Luning, L., Roeleveld, P. & Claessen, V. W. M.  Sludge minimization by disintegration at different wastewater treatment plants. Water Pract. Technol. 3 (1). STOWA  Slibdesintegratie, Stowarapport 2008–10. www. stowa.nl. STOWA  Praktijkonderzoek naar oorzaken schuimvorming in slibgistingstanks, STOWA rapport 2010–43. www.stowa.nl. STOWA a Verkenning thermische slibontsluiting, STOWA rapport 2011–W-03. www.stowa.nl. STOWA b Handboek slibgisting, STOWA rapport 2011–16. www.stowa.nl.

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STOWA  Thermische slibontsluiting, pilotonderzoek naar de mogelijkheden en randvoorwaarden, STOWA rapport 2012– 25. www.stowa.nl. Stuckey, D. C. & McCarty, P. L.  Thermochemical pretreatment of nitrogenous materials to increase methane yield. Biotechnol. Bioeng. Symp. 8, 219–233. Stuckey, D. C. & McCarty, P. L.  The effect of thermal pretreatment on the anaerobic biodegradability and toxicity of waste activated sludge. Water Res. 18 (11), 1343–1353.

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Tattersall, J. M., Koodie, A. V., Riches, S., de Mooij, H. W. & deBarbadillo, C.  Treating liquors from enhanced digestion facilities start-up, commissioning and optimization of a SHARON LTP at Whitlingham STC. 16th European Biosolids and Organic Resources Conference, 2011, Leeds, UK. van Dijk, L. & de Man, A.  TurboTec®: continue thermische slib hydrolyse voor lagere zuiveringskosten en meer energie op de rwzi., Neerslag 2010, nr 5.

First received 14 August 2013; accepted in revised form 17 February 2014. Available online 20 March 2014

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Monitoring the oxygen transfer efﬁciency of full-scale aeration systems: investigation method and experimental results Riccardo Gori, Alice Balducci, Cecilia Caretti and Claudio Lubello

ABSTRACT This paper reports the results of a series of off-gas tests aimed at monitoring the evolution of the oxygen transfer efﬁciency in an urban wastewater treatment plant (3,500 population equivalent) located in Tuscany (Italy). The tests were conducted over a 2-year period starting with the testing of the aeration system. It was found that in the absence of membrane-panel cleaning operations, the oxygen transfer efﬁciency under standard conditions in process water (αSOTE) dropped from 18 to 9.5% in 2 years. This gives rise to a 40% increase in the wastewater treatment plant annual energy costs. The on-site chemical cleaning of the diffusers allowed for an almost total recovery of the

Riccardo Gori (corresponding author) Alice Balducci Cecilia Caretti Claudio Lubello Department of Civil and Environmental Engineering, University of Florence, Via S.Marta, 3 - 50139 Florence, Italy E-mail: riccardo.gori@dicea.uniﬁ.it

transfer efﬁciency (αSOTE equal to 16%). The use of the off-gas method for monitoring the oxygen transfer efﬁciency over time is therefore essential for enabling correct planning of the cleaning operations of the diffusers and for cutting the energy consumption and operating costs of the aeration system. Key words

| aeration system, ﬁne-bubble, off-gas method, oxygen transfer, wastewater treatment

INTRODUCTION Energy saving as a method to reduce operating costs and CO2 emissions is becoming more and more important in wastewater treatment plants (WWTPs) (Libra et al. ). Since in a conventional activated sludge plant energy demand is largely dominated by the aeration (Rosso & Stenstrom ), considerable savings are possible by optimising its design and operation. At present, the fine-bubble aeration system is the most widely used in European and North American WWTPs. The use of these devices offers numerous advantages including energy savings of about 50% compared to the coarse-bubble aeration systems (Stenstrom et al. ). However, due to the chemical nature and morphology of the materials making up the membrane, they are subject to fouling and scaling which can affect their operation and reduce the benefits of their utilisation in the long term (US EPA ). The energy consumption of the aeration system depends both on the efficiency of its components (e.g. blowers), and on the characteristics of the wastewater, and it is not usually possible to predict the aeration system’s behaviour under process conditions. It is therefore necessary to perform a series of measurements to verify the aeration system’s doi: 10.2166/wst.2014.165

behaviour in operating conditions. For this purpose several methods have been developed over the years. Among the available methods, the off-gas method has proved to be an effective technique that offers numerous advantages for testing diffused air aeration systems (Redmon & Boyle ; Iranpour et al. ; Capela et al. ). In many countries, the issue of energy end-use efficiency has become so important that it is now incentivised and governed by regulations and standards. In Italy, the water sector has been classified as one of the ‘strategic’ sectors and the measures for improving energy efficiency are certified via the issuing of ‘Energy efficiency certificates’ (‘white certificates’). Some of the interventions which are recognised for the purpose of issuing the white certificates concern the aeration system and include the replacement of mechanical aerators with fine-bubble aeration systems and the installation of electronic devices (inverters) for adjusting the frequency and the power in electric motors. This article reports the results of three off-gas test campaigns conducted on a WWTP in Tuscany in order to quantify the efficiency of the oxygen transfer of the aeration system consisting of membrane panels. Particular attention

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Monitoring the oxygen transfer efﬁciency of full-scale aeration systems

has been paid to the economic savings deriving from the correct management of the aeration system. The tests were conducted within the AERE project (‘Increasing the energy efficiency of wastewater treatment plants’ financed by the Italian Ministry of the Environment and the Protection of Land and Sea), whose aim is to develop a prototype and a method for monitoring the oxygen transfer efficiency in operating conditions by means of the off-gas method.

MATERIALS AND METHODS Off-gas method The off-gas method was developed by Redmon & Boyle () and is based on a gas-phase mass balance for directly measuring the efficiency of oxygen transfer in aeration systems under process conditions (ASCE ). The mass balance is performed by making a comparison between the oxygen content in a flow of atmospheric air (reference) and in an off-gas flow withdrawn from above the process tanks. See Figure 1 for a schematic layout of the system. The off-gas test campaigns were conducted using an analyser developed by the Civil and Environmental Engineering Department of the University of Florence in collaboration with the Civil and Environmental Engineering Department of the University of California – Irvine. This instrument is equipped with a combustible chamber that makes it possible to measure the oxygen content in the gas flows, a column (h ¼ 0.255 m, d ¼ 0.025 m)

Figure 1

Functional off-gas test diagram.

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for absorbing the CO2 and the water vapour present in the gas flows, a hot wire anemometer for measuring the off-gas uptake captured by the hood, and a valve for choosing between sending the ambient air and the off-gas into the test circuit. The signals coming from the oxymeter and the anemometer are automatically acquired every 0.05 s using an acquisition data card (National Instruments, sbRIO-9632). In order to collect the oxidation tank off-gas, a rectangular PVC hood was used with a 0.53 m2 capturing surface, fitted with a hose with a 0.038 m diameter that allows for conveying the off-gas towards the analyser. The choice of instrumentation, in particular the shape and size, was dictated by the need to provide a system ensuring easy transport and handling due to numerous tests to be conducted on different plants. The performing of an off-gas test campaign foresees the positioning of the collection device in different points on the surface of the oxidation tank according to a sampling plan established depending on the geometry of the tank and the aeration system. For each position, we analysed alternatively the ambient air and the off-gas for 4 min; this was the necessary time for obtaining steady-state conditions of the output signals from instruments. The following parameters are measured during the analysis: the off-gas flow rate, the oxygen concentration in the gas phase, the dissolved oxygen (DO) concentration in the liquid phase, the atmospheric pressure, and the temperature of both the off-gas and the process wastewater. The recorded data allow the calculation of the oxygen transfer efficiency by means of Equations (1) and (2) under process

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Monitoring the oxygen transfer efﬁciency of full-scale aeration systems

conditions (OTE) and in standard conditions for both new (αSOTE) and used (αFSOTE) diffusers, respectively: Y(O2,ref ) Y(O2,O G ) 100 OTE ¼ Y(O2,ref )

(1) W

Cs,20 OTE θ(20 TW ( C)) 100 αSOTE (or αFSOTE) ¼ β Cs,T DO

(2)

where OTE: oxygen transfer efficiency in process conditions (%); Y(O2,ref ), Y(O2,O G ): molar ratio of oxygen compared to the inert aggregates in the ambient air and the off-gas (%); θ: empirical temperature correction factor (here set at 1.024) (non-dimensional); β: coefficient that takes the wastewater salinity into account (here calculated on the basis of total dissolved solids content) (non-dimensional); Cs,20 : concentration of oxygen saturation at 20 C (mg/l); Cs,T : concentration of oxygen saturation in operating conditions (mg/l); DO: concentration of dissolved oxygen in the tank (mg/l); α: ratio of process to clean water mass transfer for new diffusers (non-dimensional); αF: ratio of process to clean water mass transfer for used diffusers (non-dimensional); SOTE: oxygen transfer efficiency in standard conditions for clean water (%). In order to fit the experimental data into a broader framework of operation of the plant and the wastewater characteristics, the following parameters were also established at the same time as carrying out the monitoring of the aeration system: the concentrations of chemical oxygen demand (COD), ammonium, nitrites, nitrates and specific pollutants such as surfactants, that influence the oxygen transfer (Rosso et al. ), in both the oxidation tank influent and effluent. W

Figure 2

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Tested plant and tests conducted The experimental tests were conducted in the oxidation tank of an urban WWTP of 3,500 population equivalent (P.E.) located in Tuscany (Italy). The contribution of industrial wastewater is negligible. This plant was designed to operate as a conventional activated sludge system for carbonaceous substrate removal and nitrification. The biological compartment consists of a single oxidation tank with a rectangular shape, 20 m long and 8 m wide, with side water depth of 3.6 m. Initially, this plant was equipped with a system of disk diffusers uniformly distributed on the bottom of the tank. In 2010, due to serious air diffusion problems, the aeration system was replaced with 24 high-performance membrane panels (sized 4 m by 0.18 m), arranged in four frames each containing six extractable panels (Figure 2). Currently, the oxidation compartment is operated in alternate-cycles mode in order to have nitrification and denitrification in the same tank. The air flow is delivered in a timed manner with the time intervals periodically adjusted according to the purification process requirements. During the non-aerated periods mixing is assured by two mixers in opposite corners of the oxidation tank (Figure 2). The air flow rate supplied to the tank is not measured. Three off-gas test campaigns were carried out on this plant, one in June 2010 and the others in May and July 2012. At the time of the first test the aeration system was fed by a positive displacement blower (Robuschi, RV80, flow rate 1105 Nm3/h at 770 mbar) which was replaced before the second test by a positive displacement blower (Robuschi, RVT60, flow rate 485 Nm3/h at 500 mbar) which has a lower power. The operating parameters and incoming wastewater characteristics on the day of testing are summarised in Table 1.

Diagram of the test tank indicating the sampling locations (from 1 to 10) and diffuser mixer positions.

Table 1

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Monitoring the oxygen transfer efﬁciency of full-scale aeration systems

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Characteristics of the WWTP inﬂuent and efﬂuent

Test

COD (mg/l)

N-NHþ 4 (mg/l)

N-NO 2 (mg/l)

N-NO 3 (mg/l)

28.0

< 0.03

< 0.3

Flow rate (m3/h)

Surfactants (mg/l)

Tw a ( C) W

Inﬂuent June 2010

117

–

51.8

–

May 2012

189

28.6

< 0.03

< 0.3

6.1

49.5

–

July 2012

275

60.8

< 0.03

< 0.3

10.4

24.2

–

< 0.1

0.16

12.1

–

Efﬂuent June 2010 May 2012

1.1

0.06

16.1

0.1

–

July 2012

0.4

< 0.03

18.2

0.2

–

Wastewater temperature in the oxidation tank.

According to the data gathered during the tests, we can see that the plant is effective for both the removal of organic matter and nitrification. Nitrogen removal ranges between about 40 and 70%; such a variability can be attributed to the ratio between aerobic and anoxic periods, which is not continuously regulated on the basis of the influent characteristics. Moreover, the COD/N-NHþ 4 ratio, which can affect the nitrogen removal is variable (range of 4.2–6.6 is consistent with the long-term observations of the plant manager). The sludge age of the plant was estimated to be about 44 d, and thus the plant can be considered to be of the extended aeration type. In each test campaign the measurements were taken in 10 different locations inside the tank, as illustrated in Figure 2, so as to guarantee the most uniform coverage of the tank possible. The total area sampled complies with the minimum coverage limit of 2% recommended by the guidelines (ASCE ). It must be noted that during the period elapsing between the first and the second tests, the blower of the plant was replaced with another with a lower power. Therefore it was not possible to make the aeration system work under exactly the same conditions even though steps were taken to ensure conditions as similar as possible for making the data comparable. The first test was conducted in June 2010, immediately after the installation of the new membrane panels. The initial goal was to verify whether the new aeration system was able to guarantee the performance claimed during the sizing. The next test was conducted 2 years later, in May 2012, to see whether there had been any loss of efficiency over time especially since no cleaning operations had been carried out in the meantime. Finally, another test was conducted in July 2012 immediately after carrying out a cleaning operation of the diffusers with peracetic

acid, without extracting the panels from the tank. Results of the off-gas tests were then used to assess the aeration efﬁciency recovery due to the cleaning.

RESULTS The overall αFSOTE of the aeration system was obtained by weighting the local efficiency measured at each hood location by the gas flow rates collected in the sampling points. The average values describing the performance of the aeration system during the three campaigns are summarised in Table 2. The test conducted in June 2010 confirmed that the membrane panels are characterised by high-performance oxygen transfer and the values provided by the manufacturer and used during the sizing are representative of the system under examination. From an analysis of the situations hypothesised during the design stage, which take all the specific site characteristics into account, it was in fact found that the test could be considered positive for αSOTE values higher than or equal to 15%. After 2 years’ operation there was a large reduction in the performance of the diffusers. This reduction was quantified, via use of the off-gas method, as a halving of the αSOTE

Table 2

Results of the off-gas tests

Conditions of the

OTE

αFSOTEa

Test

diffusers

(mg/l)

(%)

αFa

June 2010

New

4.62

9.8

18.0

0.53

May 2012

2 years’ service

0.03

9.4

9.5

0.30

July 2012

Cleaned

3.70

9.2

15.8

0.46

For new diffusers αSOTE and α.

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Monitoring the oxygen transfer efﬁciency of full-scale aeration systems

registered with new diffusers. A loss of panel efficiency was conceivable after observing how the oxygen concentration in the tank was much lower than that found in 2010, and it was not possible to attribute this phenomenon to the lower air flow delivered to the system alone. Without using the off-gas method it was not possible to quantify the reduction in the transfer efficiency. It is important to highlight that the low DO concentration registered in the second test could affect the off-gas test due to simultaneous nitrification–denitrification. For this reason a nitrogen mass balance was carried out on a daily basis, and, even considering the denitrification rate constant all day long, we estimated that the amount of nitrogen produced from denitrification contributes less than 1% to the nitrogen already present in the off-gas. Consequently we neglected this contribution in the calculation of OTE. The deterioration of the fine-bubble diffuser operating performance over time, which has been widely documented in literature, is caused by the formation of bubbles with larger dimensions due to the biofilm that deposits on the membrane of the diffusers (Rosso & Stenstrom ). In this specific case it is possible that the dirtying of the diffusers was accelerated by the change of operational management of the plant, establishing an anoxic phase during which the aeration system is turned off. Following the cleaning of diffusers, an increase equal to approximately 6% was observed in the αFSOTE. A pressure drop in the air delivery pipe from 456 to 390 mbar was also noted at the same time. Table 3 shows a comparison between the off-gas flow rate and DO obtained in the same positions in different test campaigns. The flows recorded in May 2012 are systematically lower than those in 2010 due to both the different blower power between the tests and the fouling of diffusers. It can also be observed how this reduction is of the same magnitude for all the positions, and it can therefore be assumed that there

Table 3

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were no diffuser groups that were affected to a greater extent by the aging effect, or areas of the tank more prone to biological deposits. After the cleaning operation a significant increase in the off-gas flow rate was recorded in seven position out of the 10 tested. Variability in the ratio between the air flow rate before and after the cleaning can be due to the following reasons: air distribution is not uniform (not even in the first test with new diffusers), and the effect of the cleaning procedure could have been not uniform. In addition the air is supplied to the aeration tank using only one pipe and therefore fouling of diffusers can influence the air distribution. Data in Table 3 also show that the DO concentrations recorded in July 2012 were significantly higher than those registered in May 2012 for all points monitored (mean average of 3.7 and 0.03 mg/l, respectively). The results obtained therefore confirm what was already expressed for the oxygen transfer efficiency data and, namely, the positive outcome of the intervention performed. The fact of being able to quantify the increases or losses of efficiency of the aeration system plays a vital role in the economic management of a plant, since it makes it possible to quantify the increases or reductions in the energy consumption linked to the operation of the blowers, and consequently, to the operating costs of the plant. Indeed, with the same oxygen required for the biological reactions, an increase in performance in the aeration system gives rise to a reduced air flow necessary to supply the tanks, and, as a result, lower power consumption by the blowers. An economic analysis was then carried out in order to evaluate the effect of diffuser fouling on the treatment cost. We firstly considered the values set for the design of the aeration system: 34.2% for SOTE (from diffuser’s technical sheet), 0.276 kgO2/Nm3 for the oxygen content in the air flow in normal conditions (WO2), 0.55 for the αF-value, 290 Nm3/h for the air flow rate, and 0.99 and 1.024 for the coefficients β and θ, respectively. For the comparison

Comparison of off-gas ﬂow rate (OG-FR) and DO (mg/l) obtained during the tests Positions

Test

June 2010

OG-FR DO

3.97 1.43

4.02 3.83

3.90 4.80

3.97 5.43

2.78 5.67

2.84 5.81

3.31 5.88

2.75 5.31

2.77 4.71

5.44 3.38

May 2012

OG-FR DO

0.45 0.01

0.40 0.07

0.36 0.01

0.29 0.04

0.56 0.01

0.48 0.04

0.67 0.06

0.56 0.05

0.49 0.03

0.50 0.01

July 2012

OG-FR DO

0.48 1.26

0.82 3.89

0.39 4.43

0.66 4.34

0.88 4.11

0.35 1.55

1.19 3.97

0.55 4.5

0.8 4.2

0.86 4.7

OG-FR of June 2010 is expressed in Nm3/h. OG-FR of May 2012 and July 2012 is expressed as ratio to the corresponding value of June 2010.

Table 4

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Monitoring the oxygen transfer efﬁciency of full-scale aeration systems

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Blower energy consumption with variations in αFSOTE

Test

αFSOTE (%)

Air ﬂowa (Nm3/h)

Blower power (kW)b

Energy (kWh/y)

kWh kg CODrem

June 2010

18.0

301

4.8

43,800

0.68

6570

May 2012

9.5

568

8.2

83,220

1.28

12,483

July 2012

15.8

343

5.0

50,808

0.78

7621

Aeration cost (€/y)

Air ﬂow rate necessary to guarantee an OTR ¼ 11.6 kgO2/h. Power required to ensure the necessary OTR in process conditions (estimation was carried out on blower’s technical sheet).

purposes we considered the following conditions: DO 2 mg/l and temperature 20 C; thus Cs,T was set at 9.2 mg/l. The oxygen transfer rate (OTR) was then calculated using the following expression: W

αFSOTE Q WO2 θ(20 TW ( C)) (β Cs,T DO) OTR ¼ Cs,20 ¼ 11:6 kgO2=h

(3)

We then used the Equation (3) to estimate the air flow rate required to ensure the same OTR with the αFSOTE registered during the tests. The calculated flow rate and the pressure drop reduction measured after the cleaning operation were then used to estimate the power required by the blower in the different conditions, according to its technical sheet (see Table 4). Results of the economic analysis are summarised in Table 4. It can be observed how a reduction in the αFSOTE value from 18.0 to 9.5% gives rise to an increase in management costs by approximately €5,900/y. The recovery of the aeration efficiency due to the cleaning allows a significant reduction in the annual energy consumption and a saving of about 40% of the aeration cost with respect to the worst case. In terms of environmental effects, considering a specific 0.406 kgCO2/kWh as specific emission value (IEA ), the energy saving obtained with cleaning of diffusers corresponds to 13.2 tCO2/y and 3.76 kgCO2/y P.E.

CONCLUSIONS This paper shows the results of a case study on the use the off-gas method, and on its subsequent energy and economic savings. This method was applied to a plant in Tuscany (Italy) equipped with micro-perforated membrane panels, and it was found that after 2 years’ operation, the oxygen transfer efﬁciency dropped from 18 to 9.5%. After washing the diffusers with peracetic acid it was possible to recover the majority of lost

efficiency and re-establish values resembling the initial ones. It was observed how a delay in cleaning the diffusers gave rise to a relevant increase in operating costs, equal to approximately €5,900/y for a plant with 3,500 P.E., which therefore cancels all the benefits of using high-performance fine-bubble aeration systems. The economic, energy and environmental savings resulting from correct management of the aeration system are therefore high and tangible, even for small-sized plants.

ACKNOWLEDGEMENT We wish to thank the Italian Ministry of the Environment and Protection of Land and Sea for supporting this research.

REFERENCES ASCE (American Society of Civil Engineers)  Standard Guidelines for In-Process Oxygen Transfer Testing. ASCE 18-96, American Society of Civil Engineers, New York. Capela, S., Gillot, S. & Héduit, A.  Comparison of oxygentransfer measurement methods under process conditions. Water Environ. Res. 76 (2), 183–188. IEA  CO2 Emissions from Fuel Combustion. International Energy Agency, Paris. Iranpour, E., Magallanes, A., Zermeno, M., Varsh, V., Abrishamchi, A. & Stenstrom, M. K.  Assessment of aeration basin performance efficiency: sampling methods and tank coverage. Water Res. 34, 3137–3152. Libra, J. A., Schuchardt, A., Sahlmann, C., Handschag, J., Wiesmann, U. & Gnirss, R.  Comparison of efficiency of large-scale ceramic and membrane aeration system with the dynamic offgas method. Water Sci. Technol. 46 (4–5), 317–324. Redmon, D. & Boyle, W. C.  Oxygen transfer efficiency measurements in mixed liquor using off-gas techniques. J. Water Pollut. Control Fed. 55 (11), 1338–1347. Rosso, D. & Stenstrom, M. K.  Comparative economic analysis of the impacts of mean cell retention time and denitrification on aeration systems. Water Res. 39 (16), 3773–3780.

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Monitoring the oxygen transfer efﬁciency of full-scale aeration systems

Rosso, D. & Stenstrom, M. K.  Economic implications of ﬁne-pore diffuser aging. Water Environ. Res. 78 (8), 810–815. Rosso, D., Huo, D. L. & Stenstrom, M. K.  Effects of interfacial surfactant contamination on bubble gas transfer. Chem. Eng. Sci. 61, 5500–5514.

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Stenstrom, M. K., Vazirinejad, H. O. & Ng, A.  Economic evaluation of upgrading aeration system. J. Water Pollut. Control Fed. 56 (1), 20–26. US EPA  Fine Pore Aeration System. EPA/625/1-89/023, United States Environmental Protection Agency, Cincinnati, Ohio, USA.

First received 5 September 2013; accepted in revised form 17 March 2014. Available online 28 March 2014

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Inverse calculation of biochemical oxygen demand models based on time domain for the tidal Foshan River Li Er and Zeng Xiangying

ABSTRACT To simulate the variation of biochemical oxygen demand (BOD) in the tidal Foshan River, inverse calculations based on time domain are applied to the longitudinal dispersion coefﬁcient (E(x)) and BOD decay rate (K(x)) in the BOD model for the tidal Foshan River. The derivatives of the inverse calculation have been respectively established on the basis of different ﬂow directions in the tidal river. The results of this paper indicate that the calculated values of BOD based on the inverse

Li Er (corresponding author) Zeng Xiangying Wuhan Municipal Engineering Design and Research Institute, Changqing Road 40, Jianghan District, Wuhan 430015, China E-mail: tjy903@163.com

calculation developed for the tidal Foshan River match the measured ones well. According to the calibration and veriﬁcation of the inversely calculated BOD models, K(x) is more sensitive to the models than E(x) and different data sets of E(x) and K(x) hardly affect the precision of the models. Key words

| BOD decay rate, inverse calculation, longitudinal dispersion coefﬁcient, time domain

INTRODUCTION In recent years, concerns have grown on the pollution of the tidal river in China. Tides are the periodic rise and fall of the sea level due to attractive forces of the sun, moon, and earth. Tides and tidal currents are a major source of energy for turbulence and mixing in estuaries and they play an important role in the movement of dissolved and particulate material, creating oscillatory fluxes in physical and chemical properties. The dissipation of tidal energy causes changes in the vertical stability of the water column. Denman & Powell () showed that tidal mixing is important for phytoplankton and primary productivity because it produces light and nutrient fluctuations. Yin et al. () showed that the tidal-regulated estuarine circulation played an important role in entrainment of nutrients in the Fraser River estuary and adjacent coastal waters of the Strait of Georgia (Canada). Compared with a non-tidal river, pollutants in a tidal river distribute uniformly throughout the whole cross-sections more rapidly. Therefore a one-dimensional water quality model can be fit for a tidal river (Atkinson et al. ; Ghermandi et al. ; Franceschini & Tsai ; Mannina & Viviani ). Biochemical oxygen demand (BOD) is a crucial index indicating organic pollution status of water, and the study of BOD models for rivers is growing (Ani et al. ; Telci et al. ; Erturk et al. ). As crucial parameters, longitudinal dispersion coefficient (E(x)) and doi: 10.2166/wst.2014.177

BOD decay rate (K(x)) of one-dimensional water quality models for BOD vary with different downstream distance of a tidal river (Wagener et al. ). Inverse calculation is an effect mathematical method to develop a tidal river’s BOD model ( Paredes et al. ; Freni et al. ), for it has such characteristics as high efficiency and accuracy in turning infinite variables into finite equations, and fast transformation of the corresponding initial and boundary conditions of the equations into finite constrained ones (Beck et al. ; Zheng & Gore ). The objective of this paper is to establish the inverse calculation of the BOD model on the basis of E(x) and K(x) for the tidal Foshan River in China. As a feature, the increasing number of the distributed parameters in an inverse calculated model tends to lead to the occurrence of illconditioned equations (Zheng & Gore ; Erturk et al. ). In addition multi-parameter inverse calculation equations cause high nonlinear problems very easily (Beck et al. ). Therefore the conventional optimal parameters estimation methods such as discrete optimizing theory and pulse-spectrum technique are not fit for inversely calculated BOD models, and the compound approach in which inverse calculation developed based on time domain is in effect (Beck et al. ; Zheng & Gore ). The inverse calculation of BOD models in time domain for the tidal Foshan River is calibrated and verified in this work.

Figure 1

L. Er & Z. Xiangying

Inversely calculated BOD models for the tidal Foshan River

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(a) Proﬁle map of the tidal Foshan River including location of stations and (b) location of the tidal Foshan River (circle).

PROFILE OF THE FOSHAN RIVER The tidal Foshan River, connected with the tidal Nanhai River on its western and eastern estuary, is situated in southern China (shown in Figure 1). The width of the river ranges from 0.3 to 0.8 km and the length of it is about 28.5 km. The tide of this river mainly comes from the Pacific Oceanic tidal propagation through the Dongping Strait with a mean tidal range between 1.0 and 1.7 m, and also affected by the geometry and bottom topography, and meteorological factors (Er et al. ; Svejkovsky et al. ). There have been large-scale field observations and research on tides and tidal flows in the tidal Foshan River in the past, from which it can be concluded that the tidal cycle in this area is mainly a semidiurnal mixed tidal regime with daily inequality in range and time between the period of high flows and low flows. In addition the water level of the Nanhai River influences the tide of the tidal Foshan River. When its water level is less than 1.0 m, the tide appears considerably strong, and when its water level exceeds 1.7 m the tide becomes quite weak. The pattern of residual flows in the tidal Foshan River reciprocates owing to the different tidal force between its western and eastern part (Er et al. ; Svejkovsky et al. ). Hydrological monitoring and water quality monitoring stations were distributed in the whole river and its tributaries. The stations a–k (called Stn a–k for short) and the stations T1–T5 (called Stn T1–T5 for short) were set up in the main river and its tributaries for measuring their respective BOD (ultimate carbonaceous BOD (CBODu)), flow rate and flow velocity, and the stations G1–G2 (called Stn G1–G2 for short) were established in the estuary of the main river for measuring

its tidal indices. Field measurements of Stn a–k and T1–T5 were obtained during cruises in March, April, June, July, September, October, November and December in 2012, and those of Stn G1–G2 were got in the whole of 2012. Flow velocity and ﬂow direction at all the stations were consecutively recorded with a Nortek ADP-1500 mounted on the sediment bottom. The values of BOD were measured based on the standardized procedure (Ani et al. ; Xiangying et al. ).

MODEL One-dimensional BOD models for the tidal Foshan River The BOD model of the tidal Foshan River may be consecutively established on a reach by reach basis. According to the measurements of each sampling station, the maximum BOD and ﬂow rates from the tributaries of the river are shown in Figure 2. The ranges of the ratio of the ﬂow rate and total BOD of the tributaries to the main river were measured

Figure 2

BOD and ﬂow rate contributed from tributaries of the tidal Foshan River. BODmax is the maximum BOD of each tributary during the measurement period; qmax is the maximum ﬂow rate of each tributary during the measurement period.

L. Er & Z. Xiangying

Inversely calculated BOD models for the tidal Foshan River

as follows: 0.16–2.03% and 0.15–0.72% in the Luocung River, 0.21–1.93% and 0.11–0.65% in the Nanbei River, 1.1–3.21% and 0.8–1.6% in the Junqiao River, 0.22–1.88% and 0.17–0.61% in the Jiaobian River, and 0.23–2.28% and 0.26–0.79% in the Huadi River respectively. Therefore, it can be concluded that the ﬂow rate and total BOD of the main stream are more than those from each tributary. And past measurements have shown that the total BOD from other potential pollution sources along tributaries contribute little to the main stream based on past measured (Xiangying et al. ). Therefore under unsteady ﬂow conditions, the one-dimensional BOD model together with its initial and boundary conditions, ignoring the component caused by other potential pollution sources and tributaries (Ghermandi et al. ), can be developed as follows: 8 @C(x, t) @C(x, t) @ @C(x, t) > > > þu ¼ E(x) K(x) C(x, t) > > @x @x @x < @t C(0, t) ¼ f(t)C(l, t) ¼ g(t)t ∈ [0, T ]C(x, 0) > > > @C(x, t) > > ¼ φ(x)x ∈ [0, l] ¼ 0t ∈ [0, T ] : @x x¼l

(1)

where x is stream distance (L), t is stream time (T), u is crosssectional flow velocity (LT 1), C(x, t) is concentration of CBODu (ML 3) at a distance x(L) downstream at a time t(T), E(x) is longitudinal dispersion coefficient (L2T 1), K(x) is BOD decay rate at a distance x (T 1), T is the whole time of the stream time (T), @=@t is partial differential expression of t; @=@x is partial differential expression of x, l is whole length of the river (L), f(t)gðtÞ is the function of t and φ(x) is the function of x. The inverse calculation based on time domain in different flow direction The flow direction of a tidal river A tidal river can be divided into many reaches and there are four types of flow directions existing in each reach in a tidal river during a tide: forward flow, backward flow, divergent flow and convergent flow. In this paper, the flow directions from a reach’s western side to its eastern side and from a reach’s eastern side are respectively called as ‘forward flow’ and ‘backward flow’. The flow directions from a reach’s centre to its two sides and from both sides of a reach to its centre are respectively defined as ‘divergent flow’ and ‘convergent flow’. Both ‘forward flow’ and

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‘backward flow’ are called as ‘unidirectional flow’, and both ‘divergent flow’ and ‘convergent flow’ are named as ‘multidirectional flow’. Obviously there are three directions in (‘forward flow’, ‘backward flow’ and ‘multidirectional flow’) all reaches of a tidal river at the same time, and the BOD model of a tidal should be developed in these respective directions.

The inverse calculation models based on time domain Under the forward ﬂow direction condition, the numerical solution of Equation (1), together with its initial and boundary conditions, is obtained by discretizing it over a nonuniform rectangular grid and the corresponding unconditional stable implicit ﬁnite difference scheme as follows: 8 Ci,jþ1 Ci,j Ciþ1,jþ1 Ci,jþ1 > > þu > > τ h > > > > > < ¼ Eiþ1 Ciþ1,jþ1 (Eiþ1 þ Ei )Ci,jþ1 þ Ei Ci 1,jþ1 Ki Ci,jþ1 h2 > > C0,j ¼ f(jτ), Ci,0 ¼ φ(ih), CNþ1,j ¼ CN,j , CN,j ¼ g(jτ) > > > nþ1 > n n > Ci,j ¼ Ci,j þ δCi,j , Enþ1 ¼ Eni þ δEni , Kinþ1 ¼ Kin þ δKin > i > : ði ¼ 0, 1, 2, :::N, j ¼ 0, 1, 2, :::W Þ (2) where τ is the time step (T), h is the distance step (L), and n , Eni and Kin represent the nth iterative value of C(xi , tj ), Ci,j E(xi ) and K(xi ) in the ith reach respectively. xi is the ith n spatial grid stream distance, tj is the ith grid time, δCi,j is nþ1 n n the difference between Ci,j and Ci,j , δEi is the difference between Enþ1 and Eni , and δKin is the difference between i Kinþ1 and Kin . The transition matrix M of Equation (2) is expressed as follows: λ ð1 μ þ μ cos KhÞ i sin Kh 2 M¼ λ ð1 þ μ μ cos Kh þ Kτ Þ þ i sin Kh 2

(2a)

where λ ¼ u(τ=h), μ ¼ (τ=3h2 )(Ei 1 þ Ei þ Eiþ1 ), and the determinant of the matrix M is the following expression:

2 λ sin Kh 2 jM j2 ¼ 2 λ ð1 þ μ μ cos Kh þ Kτ Þ2 þ sin Kh 2 ð1 μ þ μ cos KhÞ2

(2b)

L. Er & Z. Xiangying

Inversely calculated BOD models for the tidal Foshan River

According to Equation (2b), the following equation is obtained:

jMj2 1 ¼

8μsin2 Kh Kτ(2 þ 4μ sin2 Kh þ Kτ) 2 λ (1 þ μ μ cos Kh þ Kτ)2 þ sin Kh 2

(2c)

Obviously, jMj2 1 0, therefore jMj2 1. So based on the equivalence theorem, Equation (2) is a convergence equation. According to Equation (2), the expression of the forward n calculation model about Ci,j and its initial and boundary conditions can be yielded as follows: 8 u Eniþ1 1 u Eni þ Eniþ1 > n n n > > 2 Ciþ1,jþ1 þ þ þ Ki Ci,jþ1 > > τ h h h2 < h En n 1 n > 2i Ci 1,jþ1 ¼ Ci,j > > τ h > > n n n n : C0,j ¼ f(jh), Ci,0 ¼ φ(ih), CNþ1,j ¼ CN,j

The inverse calculation model about be written as follows:

δEni

and

δKin

(3)

The Green function Gi,j,p,q is developed as follows:

2014

K δCN,q ¼

N 1 X <W 1 X i¼1

n n [Gi,j,N,q (Ci 1,jþ1 Ci,jþ1 )

j¼1

δEn n n Ci,jþ1 )] 2i þGiþ1,j,N,q 1 (Ciþ1,jþ1 h 9 W 1 = X n Gi,j,N,q Ci,jþ1 δKin ; j¼1

(7)

Equation (7) is the inverse equation for the forward flow direction, and those for other flow directions are similar to it, shown as follows. Backward flow:

K δCN,q

N 1<W 1 X X

: i¼1

n n [Gi 1,j,N,q (Ci,jþ1 Ciþ1,jþ1 )

j¼1

δEni 2 h9 W 1 = X n Gi 1,j,N,q Ciþ1,jþ1 δKin ; j¼1

n n Ciþ1,jþ1 )] þGi,j,N,q 1 (Ci,jþ1

(8)

Divergent ﬂow: 8

K δCN,q

N 1 X <W 1 X

: i¼1

n n [Gi,j,N,q (Ci,jþ1 Ci 1,jþ1 )

j¼1

δEn n n Ciþ1,jþ1 )] 2i þGiþ1,j,N,q 1 (Ci,jþ1 h 9 W 1 = X n þ Gi,j,N,q Ci,jþ1 δKin ; j¼1

(9)

Convergent ﬂow:

8 En 1 u Eni Eniþ1 > n > > Gi,jþ1,p,q þ þ K 2i Gi 1,jþ1,p,q þ > i > τ h h2 < h n u Eiþ1 1 > 2 Giþ1,jþ1,p,q Gi,j,p,q ¼ Gi,j,p,q þ > > h τ h > > : ði ¼ 0, 1, 2, :::N, j ¼ 0, 1, 2, :::W Þ

K ¼ δCN,q

N 1<W 1 X X

: i¼1

n n [Gi 1,j,N,q (Ciþ1,jþ1 Ci,jþ1 )

j¼1

δEni 2 h9 W 1 = X n þ Gi 1,j,N,q Ciþ1,jþ1 δKin ; j¼1 n n Ci,jþ1 )] þGi,j,N,q 1 (Ciþ1,jþ1

(5) where 1(i ¼ p, j ¼ q): 0(i ≠ p, j ≠ q)

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expression can be got based on Equations (4)–(6):

can

(4)

8 n δCi,j > Eni 1 u Eni þ Eniþ1 > n n > þ þ δC þ K δC > i i 1,jþ1 2 i,jþ1 > τ h τ h h2 > > > > n n > δE u Eiþ1 > n n n > 2 ¼ i 1 (Ci 1,jþ1 Ci,jþ1 ) þδCiþ1,jþ1 < h h h2 > δEn n > n n > Ciþ1,jþ1 ) δKi Ci,jþ1 þ 2i (Ci,jþ1 > > > h > > n n n n > δCi,0 ¼ δC0,j ¼ δCN,j ¼ δCNþ1,j ¼0 > > > : ði ¼ 0, 1, 2, :::N, j ¼ 0, 1, 2, :::W Þ

Gi,j,p,q ¼

Water Science & Technology

(6)

Considering the difference equation and its boundary conditions, G0,j,p,q 1 ¼ GN,j,p,q 1 ¼ 0. And the following

(10)

Calibration for inverse calculation models In this study Monte-Carlo Analysis (MCA) is used for assessing the sensitivity of the inverse calculation models. MCA

L. Er & Z. Xiangying

Inversely calculated BOD models for the tidal Foshan River

uses regional sensitivity analysis (RSA) to assess the sensitivity of model parameters (Soltani et al. ; Yoshiaki et al. ), where sensitivity is defined as the effect of changes in the parameters on the overall model performance (as indicated by an objective function). Sensitivity is only one of the essential requirements of an identifiable parameter. A parameter is termed identifiable if it is possible to determine its value with relative confidence within the feasible parameter space based on the model output produced. However, the values of sensitive parameters that produce a behavioral model output can be distributed over a range of the feasible parameter space and can change when they are estimated from different response modes (Maier et al. ). Hence MCA uses an identifiable analysis, which is based either on simple inspection of the ‘dotty plots’ of Monte-Carlo outputs for a given parameter, for example taken from the upper performance quantile, or from an automated analysis. For water quality models, calibration and prediction periods are assessed using the Nash–Sutcliffe criteria

Figure 3

Regional sensitivity plot of E(x) and K(x) in the inverse BOD model.

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(Telci et al. ): N P

R2 ¼ 1 i¼1 N P

(oi pi )2 (11) )2 (oi o

i¼1

where oi is the ith observed output variable, pi is the ith predicted output variable, N is the number of observations and the over-bar denotes the mean for the period of evaluation. Note that the Nash–Sutcliffe criterion has a basic reliance on the square of the errors, which thus focuses on ﬁtting peak values.

RESULTS AND DISCUSSION The state of the flow In this paper, based on the value of the flow rate in the tidal Foshan River, the state of the flow is divided into three grades: high flow, moderate flow and low flow. And the state is called

L. Er & Z. Xiangying

Inversely calculated BOD models for the tidal Foshan River

‘high flow’, ‘moderate flow’ and ‘low flow’ when the value of the flow rate is >60 m3/s, between 20 m3/s and 60 m3/s and <20 m3/s respectively. According to the measurements, the high flow appears in the period 11 August to 13 November 2012, while the moderate flow happens during 26 March to 10 August 2012 and 13 November to 13 December 2012, and the low flow occurs 1 January to 25 March 2012 and 14 December to 30 December 2012. The inverse BOD models are developed on the basis of the different state of the flow.

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sensitive to a particular parameter, given an a priori uniform distribution each group will plot on a straight line (Wagener et al. ). Five quantile groups are ranked from the highest to the lowest likelihood in Figure 3. The result shows that E(x) is more sensitive for the models of the multidirectional flows than those of the unidirectional flows. The similar conclusion can be drawn from the regional sensitivity plot of K(x). In addition, K(x) is more sensitive to the BOD models than E(x) under the same flow direction condition.

Calibration of the model Figure 3 shows a RSA of the behavior of the model parameters. The objective functions are then transformed into likelihood values (i.e. the chance of occurrence) split into ﬁve quantile groups, and the cumulative frequency distribution is calculated and plotted (Wagener et al. ). If the model performance is sensitive to a particular parameter there will be a large difference between the cumulative frequency distributions of the quantile groups. If the model performance is not

Figure 4

Comparison of the calculated BOD with the measured BOD.

Application of inversed BOD models to the tidal Foshan River As an example, the variations of flow rate and flow velocity with hours in Stn b, e, g and i on 25 June are studied. The results shows that the unidirectional flows tend to occur when flow rate is >20 m3/s and flow velocity is >40 cm/s, and the multidirectional flows happen while flow rate is <20 m3/s and flow velocity is <0.4 cm/s.

L. Er & Z. Xiangying

Inversely calculated BOD models for the tidal Foshan River

To examine the representativeness of the measured BOD obtained from central points of a cross-section, the measured BOD in the different sampling points of the cross-section of Stn b, e, g and i are measured as an example. The result shows that the range of the relative difference

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between the measured BOD from the central points and those from the other different points of each cross-section is 1.83 to 4.28%. Therefore the measured BOD from central points with 95% conﬁdence limits can be a representative for all the measured ones from other points

Figure 5

Probability of E(x) in different ﬂow directions.

Figure 6

Calculated BOD vs measured BOD with the different data sets ((a)–(c) for E(x), (d)–(f) for K(x)).

L. Er & Z. Xiangying

Inversely calculated BOD models for the tidal Foshan River

of the whole cross-section. Therefore in this work, the following measured BOD used for study are all from the central point of each cross-section. Comparison of the inversely calculated BOD with the measured ones is shown in Figure 4. The minimum, maximum and average of relative difference in the forward flow are 1.65%, 2.63% and 2.15% respectively; the minimum, maximum and average one in the backward flow are 1.48%, 2.95% and 2.51% respectively; and the minimum, maximum and average one in the multidirectional flow are 2.17%, 3.19% and 2.66% respectively. Therefore the match between the inversely calculated BOD and the measured ones is satisfactory. Further verification of the parameters of the models To further verify the validity of the models, the performance of the models is studied under different data sets for the estimations of E(x) and K(x). Figure 5 shows the statistical probability of the calculated E(x) and K(x) in high flow, moderate flow and low flow. It can be concluded from Figure 5 that most of the values of E(x) and K(x) are within 3–4 km2/h and 0.5–0.8 d 1 respectively. Therefore three data sets such as even-determined, over-determined, underdetermined one are used, and in this paper the data set of E(x) is defined as ‘under-determined’, ‘even-determined’ and ‘over-determined’ measurement when the value of E(x) is <3 km2/h, 3–4 km2/h and >4 km2/h respectively. And the data set of K(x) is named as ‘under-determined’, ‘evendetermined’ and ‘over-determined’ when the value of K(x) is <0.5 d 1, 0.5–0.8 d 1 and >0.8 d 1 respectively. Figure 6 shows the values of measured BOD as a function of the calculated values from the models in the above three different data sets for E(x) and K(x) under different flow directions condition. The results show that all of the regression coefficients between the calculated and measured values are >0.96, and the ranges of relative differences between the calculated and measured ones are all <5.1%. Thereby the calculated results match the measured ones well. This indicates that the different data sets hardly affect the performance of the inversely calculated models under different flow directions condition.

CONCLUSION The inverse calculations of longitudinal dispersion coefﬁcient (E(x)) and BOD decay rate (K(x)) of the BOD models are developed based on time domain under different

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ﬂow directions conditions for the tidal Foshan River and their steady solutions are established. The results demonstrate that the calculated values of BOD match the measured ones well and this match is affected scarcely by different data sets (over-determined, even-determined, under-determined) for the estimation of E(x) and K(x).

REFERENCES Ani, E.-C., Wallis, S., Kraslawski, A. & Agachi, P. S.  Development, calibration and evaluation of two mathematical models for pollutant transport in a small river. Environmental Modelling & Software 24 (10), 1139–1152. Atkinson, S. F., Johnson, D. R., Venables, B. J., Slye, J. L., Kennedy, J. R., Dyer, S. D., Price, B. B., Ciarlo, M., Stanton, K., Sanderson, H. & Nielsen, A.  Use of watershed factors to predict consumer surfactant risk, water quality, and habitat quality in the upper Trinity River, Texas. Science of The Total Environment 407 (13), 4028–4037. Beck, J. V., Blackwell, B. & Haji-Sheikh, A.  Comparison of some inverse heat conduction methods using experimental data. International Journal of Heat and Mass Transfer 39 (17), 36–49. Denman, K. L. & Powell, T. M.  Effects of physical processes on planktonic ecosystems in the coastal ocean. Oceanography and Marine Biology 22 (4), 125–168. Er, L., Yuehua, F. & Man, L.  The Evolution and Models of the Foshan River Estuary. Huazhong University of Science and Technology Press, Wuhan, China, pp. 33–72 (in Chinese). Erturk, A., Gurel, M., Ekdal, A., Tavsan, C., Ugurluoglu, A., Seker, D. Z., Tanik, A. & Ozturk, I.  Water quality assessment and meta model development in Melen watershed – Turkey. Journal of Environmental Management 91 (7), 1526–1545. Franceschini, S. & Tsai, C. W.  Assessment of uncertainty sources in water quality modeling in the Niagara River. Advances in Water Resources 33 (4), 493–503. Freni, G., Mannina, G. & Viviani, G.  Assessment of the integrated urban water quality model complexity through identiﬁability analysis. Water Research 45 (1), 37–50. Ghermandi, A., Vandenberghe, V., Benedetti, L., Bauwens, W. & Vanrolleghem, P. A.  Model-based assessment of shading effect by riparian vegetation on river water quality. Ecological Engineering 35 (1), 92–104. Maier, H. R., Jain, A., Dandy, G. C. & Sudheer, K. P.  Methods used for the development of neural networks for the prediction of water resource variables in river systems: ﬂow status and future directions. Environmental Modelling & Software 25 (8), 891–909. Mannina, G. & Viviani, G.  Water quality modelling for ephemeral rivers: Model development and parameter assessment. Journal of Hydrology 393 (3), 186–196. Paredes, J., Andreu, J. & Solera, A.  A decision support system for water quality issues in the Manzanares River (Madrid, Spain). Science of The Total Environment 408 (12), 2576– 2589.

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Inversely calculated BOD models for the tidal Foshan River

Soltani, F., Kerachian, R. & Shirangic, E.  Developing operating rules for reservoirs considering the water quality issues: application of ANFIS-based surrogate models. Expert Systems with Applications 37 (9), 6639–6645. Svejkovsky, J., Nezlin, N. P., Mustain, N. M. & Kum, J. B.  Tracking stormwater discharge plumes and water quality of the Tijuana River with multispectral aerial imagery. Estuarine. Coastal and Shelf Science 87 (3), 387–398. Telci, I. T., Nam, K., Guan, J. & Aral, M. M.  Optimal water quality monitoring network design for river systems. Journal of Environmental Management 90 (10), 2987–2998. Wagener, T., Boyle, D. P., Lees, M. J., Wheater, H. S., Gupta, H. V. & Sorooshian, S.  A framework for the development and application of hydrological models. Hydrology and Earth System Sciences 5 (1), 100–140.

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Xiangying, Z., Er, L. & Yuehua, F.  Researches on the Polluted Material of the Tidal Foshan River and its Tributary. Press of Environment Protection Research Institute of Foshan, Foshan, China, pp. 30–357 (in Chinese). Yin, K., Harrison, P. J., Pond, S. & Beamish, R. J.  Entrainment of nitrate in the Fraser River plume and its biological implications. II. Effects of spring vs neap tides and river discharge. Estuarine Coastal and Shelf Science 40 (1), 529–544. Yoshiaki, T., Masato, F., Yasuo, M., Kouki, M. & Minoru, Y.  Natural puriﬁcation effects in the river in consideration with domestic wastewater pollutant discharge reduction effects. Journal of Environmental Sciences 22 (6), 892–897. Zheng, Y. & Gore, J. P.  Measurements and inverse calculations of spectral radiation intensities of a turbulent ethylene/air jet ﬂame. Proceedings of the Combustion Institute 30 (5), 727–734.

First received 16 August 2013; accepted in revised form 25 March 2014. Available online 16 April 2014

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Removal of methylene blue, a basic dye, from aqueous solutions using nano-zerovalent iron Simin Arabi and Mahmoud Reza Sohrabi

ABSTRACT In this research, the preparation of nanoparticles of Fe0 (nano-zerovalent iron, NZVI) as adsorbent is discussed and the capability of adsorbing methylene blue (MB) is studied. The morphology of the adsorbent was evaluated with transmission electron microscopy. Batch studies were performed to delineate the inﬂuence of various experimental parameters such as pH, adsorbent dosage, initial dye

Simin Arabi (corresponding author) Mahmoud Reza Sohrabi Department of Chemistry, North Tehran Branch, Islamic Azad University, Tehran, Iran E-mail: siminarabi1354@yahoo.com

concentration, temperature and contact time. Optimum conditions for MB removal were found to be pH 9.5, adsorbent dosage of 0.5 g L 1 and equilibrium time of 1 min. The experimental equilibrium data were adjusted by the adsorption isotherms from Langmuir and Freundlich models, and their equilibrium parameters were determined. The adsorption of MB dye by NZVI obeyed both the Freundlich and Langmuir isotherm. The adsorption capacity of NZVI for MB in terms of monolayer adsorption was 208.33 mg g 1. Key words

| adsorption, adsorption isotherm, dye removal, methylene blue, nano-zerovalent iron

INTRODUCTION Dyes have long been used in dyeing, paper and pulp, textiles, plastics, leather, cosmetics and food industries (Gulnaz et al. ). Colored material discharged from these industries poses certain hazards and environmental problems. These colored compounds are not only aesthetically displeasing but also inhibiting sunlight penetration into the stream and affecting aquatic ecosystems. Dyes usually have complex aromatic molecular structures which make them more stable and difﬁcult to biodegrade (Barragán et al. ). Furthermore, many dyes are toxic to some microorganisms and may cause direct destruction or inhibition of their catalytic capabilities. These materials are the complicated organic compounds and they resist light, washing and microbial invasions. Thus, they cannot be decomposed easily (Wang et al. ). Direct discharge of dye-containing efﬂuents into the municipal environment may cause the formation of toxic carcinogenic breakdown products. Methylene blue (MB) is a heterocyclic aromatic chemical compound with the molecular formula C16H18N3SCl. It is the most frequently used substance for dyeing silk, wood and cotton and has a number of biological uses. It has severe impact on human health; for example ingestion through the mouth may cause nausea, vomiting, diarrhea, etc. Thus, removal of MB from wastewater is of great concern from doi: 10.2166/wst.2014.189

an environmental point of view. The conventional methods for treating dye-containing wastewaters are electrochemical treatment, coagulation and ﬂocculation, chemical oxidation, liquid–liquid extraction and adsorption (Oguz & Keskinler ). Use of clean methods of low-priced and biodegradable adsorbents could be a good tool to minimize the environmental impact caused by dye manufacturing and textile efﬂuents. Currently, adsorption processes have been studied because of their low cost, easy access and effective dye removal whereby dissolved dye compounds attach themselves to the surface of adsorbents, and adsorption techniques are by far the most versatile and widely used. The most common adsorbent materials are alumina silica, metal hydroxides and activated carbon. The zerovalent iron (ZVI), an environmentally friendly strong reducing agent, has been widely studied for environmental remediation in recent years. Now, it has been used in treating and remedying wastewaters contaminated with chlorinated compounds, nitro aromatic compounds, nitrates, and heavy metals, and even for the deoxidization of more complex anthropogenic chemicals including pesticides and dyes (Uzum et al. ). Among these compounds, dyes are easier to be deoxidized than to be oxidized (Ghauch et al. ). Thus, deoxidization with ZVI is one of the most

S. Arabi & M. R. Sohrabi

The removal of MB by NZVI

important paths of degradation. An innovative method using nano-zerovalent iron (NZVI) particles to increase the specific surface area of the ZVI has also been developed for decolorization of dyes. The advantage of ZVI particles for dye color removal include ease in use as a pre-treatment process and easy recycling of the spent iron powder by removal with magnets, as well as low residual iron concentration and no necessity for further treatment of the effluents. These ZVI particles were proposed as a reactive material in permeable reactive barriers due to their great ability in reducing and stabilizing different types of pollutants (Cho & Park ). ZVI particles have many advantages, such as nontoxicity, iron being abundant in nature, low cost, high reactivity, tiny particle size and higher density of reactive surface sites. This material with particle size at the nanoscale exhibits superior activity because of its larger surface area and higher activity (Cundy et al. ). However, similar to other nanomaterials, this ultra-fine powder has a strong tendency to agglomerate into larger particles, resulting in an adverse effect on both effective surface area and catalyst performance. Palygorskite as a supporting material for NZVI was investigated (Frost et al. ). ZVI was used for MB removal and, comparing with ZVI and unmodified polygorskite, palygorskite-supported ZVI showed much higher efficiency for MB removal with a very small amount of NZVI supported. It is likely that co-precipitation with in situ generated corrosion products is the main reason for the uptake of MB from solution (Schwertmaann et al. ). The discoloration of MB in Fe0/H2O systems was investigated and the results showed that co-precipitation in Fe0/ H2O systems may be primarily regarded as a non-specific mechanism in the decolorization MB. The mechanism of MB removal by Fe0-based materials which may be suitable for environmental remediation (cast iron, low alloy steel) has not yet been systematically investigated. The mechanism of adsorption of MB and its congeners onto stainless steel particles was investigated (Imamura et al. ). MB has also been used for corrosion inhibition of mild steel in acid solution (Oguzie ). The literature on ‘Fe0 Technology’ is characterized by the fact that, since the effectiveness of Fe0 reactive walls to degrade solvents has been demonstrated, applying Fe0 to treat other compounds can be performed without previous systematic investigations. Giles and co-workers have clearly shown that the concept of contaminant adsorption and co-precipitation as fundamental removal mechanism (Giles et al. ). Miyajima and Noubactep investigated the effect of mixing Fe0 and sand on the efficiency of MB discoloration. The efficiency

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of the Fe0/sand system for contaminant reduction corroborates the adsorption co-precipitation concept (Noubactep , a, b). In this paper, the possibility of using nanoparticles of ZVI to remove MB from aqueous solution was investigated using batch adsorption studies. The effect of various factors such as contact time, pH, initial dye concentration, temperature and NZVI dosage on the removal efﬁciency of NZVI was also studied. Adsorption isotherms and kinetic studies were also investigated in order to understand the adsorption mechanism and efﬁciencies of NZVI.

EXPERIMENTAL Materials The following chemicals were used in this study. Sodium borohydride (NaBH4) and ferric chloride (FeCl3.6H2O) were purchased from Merck. MB (C.I. 52015) was obtained from Alvan Sabet Co. and was used without further puriﬁcation. All other reagents were analytical reagent grade. Deionized water was used throughout this study. Adjustment of pH of the dye solution prior to adsorption was carried out with NaOH or HCl from Merck. Preparation of NZVI particles NZVI particles were prepared by liquid phase reduction method. All solvents were degassed and saturated for 30 min with N2 before use. NZVI was synthesized by adding 1 M NaBH4 solution into 0.5 M FeCl3 solution during vigorous stirring under N2 atmosphere. The mixture’s color turned from red brown to light yellow and then eventually to black. At the same time the mixture gradually produced more black particles in the three-neck ﬂask. Ferric iron (Fe3þ) was reduced to elemental iron according to the following reaction: 4Feþ3 þ 3BH4 þ 9H2 O ! 4Fe0 ðsÞ þ 3H2 BO3 ðaqÞ þ 12Hþ þ 6H2

(1)

Then black NZVI particles were vacuum-ﬁltered and washed with deionized water and 1:1 (V/V) ethanol/ acetone. Doing so prevented the NZVI from oxidizing, and then the resulting gray-black solid was dried under nitrogen atmosphere before use (Zhang ). In addition, the surface morphology of NZVI was observed with transmission electron microscopy (TEM, JEM-2100F HR, 200 kV).

S. Arabi & M. R. Sohrabi

The removal of MB by NZVI

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Adsorption experiments

Characterization

Adsorption of MB on NZVI was carried out using a batch experiment method. The effect of contact time, solution pH, temperature and NZVI dosage was investigated. For the determination of adsorption isotherms, 50 mL of dye solution of known initial concentration (10–100 mg L 1) was shaken with a certain amount of the adsorbent (0.5 g) on a rotary shaker at 150 rpm. The contact time was varied from 0 to 140 s. Sample solutions were withdrawn at predetermined time intervals for the color removal analysis. In the pH study, the pH of MB solution was adjusted in the range of 2.5–12 by adding 0.5 M HCl or 0.5 M NaOH. About 0.5 g NZVI was then added to the solution and shaken for a predetermined time. In the experiment to investigate the effect of NZVI dosage on MB adsorption, various amounts of adsorbent in the range of 0.1–0.9 g were added to 30 mg L 1 MB without pH adjustment and shaken until equilibrium. The influence of temperature on the dye adsorption was investigated at 298, 308, 318 and 328 K under the optimized conditions of 60 s of contact time and 0.5 g of NZVI. After shaking the flasks for predetermined time intervals, all samples were withdrawn from the conical flasks and the MB solutions were separated from the adsorbent by filtration then followed by centrifugation. The residual dye concentration in the supernatant solutions was analyzed using a spectrophotometer (Varian CARY-300 UV-visible spectrophotometer, BioTech) at a wavelength of 665 nm. The adsorption capacity of dye on adsorbent was calculated using the following equation:

Figure 1(a) shows the TEM image of freshly synthesized iron nanoparticles. Surface morphology shows that there exist two layers in the NZVI particle. The layer of the inside core represents the Fe0, and the outer layer surrounding the Fe0 was iron oxide(s). Figure 1(b) shows that the MB molecules covered the NZVI surface conﬁrm the successful attachment of MB to the NZVI.

qe ¼

ðC0 Ce ÞV W

(2)

where C0 and Ce are the dye concentrations in mg L 1 initially and at a given time t, respectively, V is the volume of dye solutions in L, and w is the weight of sorbent in g.

Figure 1

TEM images of (a) NZVI and (b) NZVI after adsorption of MB.

RESULTS AND DISCUSSION Characterization of the synthesized NZVI particles The Brunauer–Emmett–Teller surface area analysis indicated that the average speciﬁc surface of the synthesized NZVI was 35 m2 g 1. Figure 1(a) shows the image of the synthesized NZVI particles as observed by TEM. The TEM image shows that most of the NZVI particles were less than 100 nm in diameter. The synthesized nanoparticles had a uniform shape and the powdery particles had a spherical morphology. The chain-like structures may have resulted from both the effects of static magnetism and surface tension. Several large-sized particles may have come from agglomeration of nanoparticles due to their magnetic properties.

Effect of contact time The inﬂuence of contact time on adsorption of MB was studied under the following experimental conditions: initial concentration 30 mg L 1, and NZVI dosage 0.5 g per 50 cc. As shown in Figure 2, uptake of dye increased instantly and reached a high value within 60 sec. This meant that a large number of adsorption sites on the NZVI surface were available at this process. Thereafter, the amount of adsorbed MB

Figure 2

S. Arabi & M. R. Sohrabi

The removal of MB by NZVI

Water Science & Technology

did not significantly change with time due to the equilibrium of the repulsive forces between the MB molecules on the solid and in the bulk phases. This trend indicated the NZVI was saturated with the adsorbate at this stage. Based on the results, the adsorption time was fixed at 1 min for the rest of the batch experiments to make sure that adsorption equilibrium was reached. Effect of initial concentration The effect of initial dye concentration on the removal of MB by NZVI is shown in Figure 3. It is evident from the figure that an increase in the initial concentration leads to a decrease in the adsorption efficiency. This is due to increase in initial dye concentration, and adsorption sites on the surface area of the NZVI are saturated. Therefore, it can be said that the adsorption is increased instantly at initial steps due to rapid attachment of dye to the surface of the NZVI, and then keeps increasing gradually until the equilibrium is reached and then remains constant. Effect of pH Dependence of dye adsorption on pH is shown in Figure 4. Dye adsorption efficiency is affected by pH variation. The adsorption of MB increased with an increase in pH. The

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Effect of contact time on the adsorption of MB onto NZVI. Figure 4

Figure 3

Effect of the initial dye concentration on the adsorption of MB onto NZVI.

Effect of solution pH on the adsorption of MB onto NZVI.

optimum pH for the adsorption of MB was found to be 9.5. This can be explained by considering the electrostatic attraction that exists between the negatively charged surface of the NZVI and MB, a cationic dye. Lower adsorption at acid pH was probably due to the presence of excess Hþ ions competing with the dye cations for adsorption site. The surface of the NZVI is positively charged and, as MB dye is a cationic dye, a positively charged surface site on the NZVI does not favor the adsorption of dye cations due to the electrostatic repulsion. At alkaline pH, the number of positively charged sites decreases and the number of negatively charged sites increases, which favors the removal of the cationic dye. Similar trends were observed for the adsorption of Malachite green onto agro-industry waste (Garg et al. ) and MB onto various carbons (Kannan & Sundaram ). Effect of NZVI dosage The effect of NZVI dosage on the removal of MB by NZVI at initial concentration (C0 ¼ 30 mg L 1) is shown in Figure 5. Results express that MB removal increase up to a certain limit and then it remains almost constant. It can be interpreted that the dosage of NZVI increased the adsorptive

Figure 5

Effect of adsorbent dosage on the adsorption of MB onto NZVI.

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The removal of MB by NZVI

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molecules into adsorbent at higher temperatures (Karthikeyan et al. ).

surface area and the number of active sites correspondingly increased since the removal process had occurred at the Fe0–H2O interface. The equilibrium removal time of NZVI in removing the MB was prolonged when the dosage of NZVI increased for a given initial concentration. This may be attributed to the decrease in removal surface area and active sites available to dye molecules resulting from intermolecular competition and formation of iron oxide on the surface of NZVI particles. Furthermore, the removed amount of dye gradually decreased as reaction time increased. A similar result has been reported for degradation of azo dye acid Black 24 when using NZVI (Shu et al. ).

Adsorption isotherm Langmuir isotherm Langmuir theory is based on the assumption that adsorption is a type of chemical combination or process and the adsorbed layer is unimolecular. The theory can be represented by the following linear form:

Effect of temperature

Ce 1 Ce ¼ þ qe Q0 b Q0

The effect of temperature on adsorption of MB on NZVI was investigated by varying adsorption temperatures at 298, 308, 318 and 328 K using 0.5 g of adsorbent NZVI at pH 9.5. Figure 6 shows that an increase in the temperature from 298 to 328 K leads to an increase in the adsorption capacity at an equilibrium time, indicating the endothermic nature of the adsorption reaction. This increases the amount of material to be adsorbed on the surface. Increase in temperature also decreases the activation energy barrier thereby increasing the rate of adsorption (Shukla et al. ). The enhancement in the adsorption capacity might be due to the chemical interaction between adsorbate and NZVI, creation of some new adsorption sites or the increased rate of intraparticle diffusion of MB

where Ce is the equilibrium concentration (mg L 1), qe is the amount of dye adsorbed at equilibrium (mg g 1) and Q0 (mg g 1) and b (L mg 1) are Langmuir constants related to adsorption capacity and energy of adsorption, respectively. The values of Q0 and b were determined for all adsorbents from intercept and slopes of the linear plots of Ce/qe versus Ce (Figure 7). The good ﬁt of the experimental data and the correlation coefﬁcient (R 2) of 0.99 indicated the applicability of the Langmuir isotherm model. The constants b and Q0 can be calculated from the slope and intercept of the plot, and the values are tabulated in Table 1. The maximum monolayer adsorption capacity of NZVI was found to be 208.33 mg g 1. The shape of the

Figure 6

Table 1

Effect of temperature on the adsorption of MB onto NZVI.

Figure 7

(3)

Langmuir adsorption isotherm for MB adsorption on NZVI.

Parameters of Langmuir and Freundlich isotherms for MB adsorption on NZVI

Langmuir isotherm Q0 (mg g 1)

b(L mg 1)

Freundlich isotherm Kf (mg1 1/n L1/n g 1)

1/n

208.33

0.6149

0.05

0.9927

3.3304

1.0352

0.9902

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Langmuir isotherm was investigated by the dimensionless constant separation factor or equilibrium parameter (RL) to determine high afﬁnity adsorption. RL was calculated as follows: RL ¼

1 1 þ bC0

(4)

where C0 is the initial dye concentration (mg L 1). RL indicates the type of isotherm to be irreversible (RL ¼ 0), favorable (0 < RL < 1), linear (RL ¼ 1) or unfavorable (RL > 1). In the present study, the value of RL was observed to be in the range 0–1 indicating the adsorption process is favorable (Figure 7).

Freundlich isotherm The Freundlich adsorption model stipulates that the ratio of solute adsorbed to the solute concentration is a function of the solution. The empirical model was shown to be consistent with an exponential distribution of active centers, characteristic of heterogeneous surfaces. The amount of solute adsorbed, qe, is related to the equilibrium concentration of solute in solution, Ce, following:

Figure 8

Freundlich adsorption isotherm for MB adsorption on NZVI.

Kinetic studies A quantitative understanding of the adsorption is possible with the help of kinetic models. The pseudo-ﬁrst-order kinetic model can be written as: log (qe qt ) ¼ log qe

k1 t 2:303

(7)

This expression can be linearized to give the following equation:

where qe and qt are the amounts of MB (mg g 1) at equilibrium and at time t, respectively, and k1 is the pseudo-first-order equilibrium rate constant (sec 1). A plot of log (qe–qt) vs t (Figure 9) gives a straight line confirming the possibility of the applicability of the pseudo-first-order rate equation (R 2 ¼ 98.71). From the plot, k1 and qe are determined from the slope and intercept respectively as shown in Table 2. The pseudo-second-order adsorption rate equation may be expressed as follows (Bouhamed et al. ):

1 log qe ¼ log KF þ log Ce n

t 1 t ¼ þ qt k2 q2e qe

qe ¼ KF Ce 1=n

(5)

(6)

where KF (L g 1) is a constant for the system related to the bonding energy. KF can be deﬁned as the adsorption or distribution coefﬁcient and represents the quantity of dye adsorbed onto NZVI for a unit equilibrium concentration (a measure of adsorption capacity, mg g 1). The slope 1/n, ranging between 0 and 1, is a measure of adsorption intensity or surface heterogeneity, becoming more heterogeneous as its value gets closer to zero. A value for 1/n below 1 indicates a normal Freundlich isotherm while 1/n above 1 is indicative of cooperative adsorption. The model parameters and R 2 value (0.99) indicated that this model showed high correlation with the experimental data (Figure 8). The value of slope 1/n (1.0352) indicates a cooperative isotherm. The adsorption of MB dye by NZVI obeyed both the Freundlich and Langmuir isotherm.

(8)

where k2 (g mg 1 s 1) is the rate constant of second-order adsorption. If second-order kinetics is applicable, the plot of t /qt versus t should show a linear relationship (Figure 10).

Figure 9

The pseudo-ﬁrst-order kinetic of MB adsorption on NZVI at 30 C.

Table 2

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The removal of MB by NZVI

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2014

The values of parameters and correlation coefﬁcients of kinetic models for MB adsorption on NZVI

Pseudo-ﬁrst-order k1(1/sec)

qe(mg g 1)

Pseudo-second-order k2(g/mg sec)

qe(mg g 1)

0.01960

21.468

0.9871

0.0022

27.472

0.9948

REFERENCES

Figure 10

The pseudo-second-order kinetic of MB adsorption on NZVI at 30 C.

The values of qe and k2 can be determined from the slope and intercept of the plot. The correlation coefficient for the pseudo-first-order ranged between 0.98 and 0.99 whereas the value for the second-order ranged between 0.99 and 1. The higher correlation coefficient indicates that the experimental data best fitted into the pseudosecond-order and suggests that the process of adsorption follows pseudo-second-order kinetics.

CONCLUSIONS The findings suggest the NZVI is a strong reducing agent for the treatment of dyeing industry wastewater. The experimental data were well described with Langmuir and Freundlich isotherm models. The pseudo-second-order kinetic model fits very well with the dynamical adsorption behavior of MB dye. It was shown that the rate of adsorption of MB on NZVI increases with the increase in temperature, thus suggesting the reaction to be spontaneous and endothermic in nature. The present study indicated that NZVI could be employed as a reactive medium for the removal of MB from industrial and other effluents.

ACKNOWLEDGEMENT The authors gratefully acknowledge Islamic Azad University, North Tehran Branch, for support.

Barragán, B. E., Costa, C. & Carmen Márquez, M.  Biodegradation of azo dyes by bacteria inoculated on solid media. Dyes and Pigments 75, 73–81. Bouhamed, F., Elouear, Z. & Bouzid, J.  Adsorptive removal of copper (II) from aqueous solutions on activated carbon prepared from Tunisian date stones: equilibrium, kinetics and thermodynamics. Journal of the Taiwan Institute of Chemical Engineers 43, 741–749. Cho, H.-H. & Park, J.-W.  Sorption and reduction of tetrachloroethylene with zero valent iron and amphiphilic molecules. Chemosphere 64, 1047–1052. Cundy, A. B., Hopkinson, L. & Whitby, L. D.  Use of ironbased technologies in contaminated land and groundwater remediation: a review. Science of the Total Environment 400 (1–3), 42–51. Frost, R. L., Xi, Y. & He, H.  Synthesis, characterization of palygorskite supported zero-valent iron and its application for methylene blue adsorption. Journal of Colloid and Interface Science 341, 153–161. Garg, V. K., Amita, M., Kumar, R. & Gupta, R.  Basic dye (methylene blue) removal from simulated waste water by adsorption using Indian Rosewood sawdust: a timber industry waste. Dyes and Pigments 63, 243–250. Ghauch, A., Tuqan, A. & Assi, H. A.  Antibiotic removal from water: elimination of amoxicillin and ampicillin by microscale and nanoscale iron particles. Environmental Pollution 157, 1626–1635. Giles, D. E., Mohapatra, M., Issa, T. B., Anand, S. & Singh, P.  Iron and aluminium based adsorption strategies for removing arsenic from water. Journal of Environmental Management 92, 3011–3022. Gulnaz, O., Kaya, A., Matyar, F. & Arikan, B.  Removal of methylene blue from aqueous solutions by graphene oxide. Journal of Hazardous Materials 108, 183–188. Imamura, K., Ikeda, E., Nagayasu, T., Sakiyama, T. & Nakanishi, K.  Adsorption behavior of methylene blue and its congeners on a stainless steel surface. Journal of Colloid and Interface Science 245 (1), 50–57. Kannan, N. & Sundaram, M. M.  Kinetics and mechanism of removal of methylene blue by adsorption on various carbons – a comparative study. Dyes and Pigments 51, 25–40. Karthikeyan, T., Rajgopal, S. & Miranda, L. R.  Chromium (VI) adsorption from aqueous solution by Hevea brasiliensis sawdust activated carbon. Journal of Hazardous Material B124, 192–199.

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The removal of MB by NZVI

Noubactep, C.  A fundamental mechanism of aqueous contaminant removal by metallic iron. Water South Africa 36, 663–670. Noubactep, C. a Aqueous contaminant removal by metallic iron: is the paradigm shifting? Water South Africa 37, 419–425. Noubactep, C. b Metallic iron for water treatment. A knowledge system challenges mainstream science. Fresenius Environmental Bulletin 20, 2632–2637. Oguz, E. & Keskinler, B.  Comparison among O3, PAC adsorption O3/HCO 3 , O3/H2O2 and O3/PAC processes for the removal of Bomaplex Red CR-L-dye from aqueous solutions. Dyes and Pigments 74, 329–334. Oguzie, E. E.  Corrosion inhibition of mild steel in hydrochloric acid solution by methylene blue dye. Materials Letters 59, 1076–1079. Schwertmaann, U., Wagner, F. & Knicker, H.  Ferrihydritehumic associations: magnetic hyperﬁne interactions. Soil Science Society of America Journal 69, 1009–1015.

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Shu, H.-Y., Chang, M.-C., Yu, H.-H. & Chen, W.-H.  Reduction of an azo dye Acid Black 24 solution using synthesized nanoscale zero valent iron particles. Journal of Colloid and Interface Science 314, 89–97. Shukla, A., Zhang, Y. H., Dubey, P., Margrave, J. L. & Shukla, S. S.  The role of sawdust in the removal of unwanted materials from water. Journal of Hazardous Materials B95, 137–152. Uzum, C., Shahwan, T., Eroglu, A. E., Lieberwirth, I., Scott, T. B. & Hallam, K. R.  Application of zero-valent iron nanoparticles for the removal of aqueous Coþ2 ions under various experimental conditions. Chemical Engineering Journal 144, 213–220. Wang, L., Zhang, J. & Wang, A.  Removal of methylene blue from aqueous solution using chitosan-g-poly (acrylic acid)/ montmorillonite super adsorbent nanocomposite. Colloids and Surfaces A: Physicochem. Eng. Aspects 32, 47–53. Zhang, W. X.  Nanoscale iron particles for environmental remediation: an overview. Journal of Nanopart Res. 5, 323–332.

First received 9 September 2013; accepted in revised form 7 April 2014. Available online 19 April 2014

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Evaluation of barrier materials for removing pollutants from groundwater rich in natural organic matter I. Kozyatnyk, P. Haglund, L. Lövgren, M. Tysklind, A. Gustafsson and N. Törneman

ABSTRACT Permeable barriers are used for passive remediation of groundwater and can be constructed from a range of materials. The optimal material depends on the types of contaminants and physicochemical parameters present at the site, as well as the hydraulic conductivity, environmental safety, availability, cost and long-term stability of the material itself. The aim of the presented study was to test a number of materials for their ability to remove heavy metals and organic pollutants from groundwater with a high (140 mg L 1) content of natural organic matter (NOM). The following materials were included in the study: sand, peat, fly ash, iron powder, lignin and combinations thereof. Polluted water was fed into glass columns loaded with each sorbent and the contaminant removal efficiency of the material was evaluated through chemical analysis of the percolate. Materials based on fly ash and zero-valent iron were found to be the most effective for heavy metal removal, while fly ash and peat were the most effective for removing aliphatic compounds. Filtration through lignin and peat led to leaching of NOM. Although the leaching decreased over time, it remained high throughout the experiments. The results indicate that remediation of contaminated

I. Kozyatnyk (corresponding author) P. Haglund L. Lövgren M. Tysklind Department of Chemistry, Umeå University, S-901 87 Umeå, Sweden E-mail: ivan.kozyatnyk@chem.umu.se A. Gustafsson MoRe Research AB, Box 70, 891 22 Örnsköldsvik, Sweden N. Törneman SWECO, Hans Michelsensgatan 2, Box 286, 201 22 Malmö, Sweden

land at disused industrial sites is a complex task that often requires the use of mixed materials or a minimum of two sequential barriers. Key words

| adsorption, groundwater, heavy metals, natural organic matter, organic pollutants, permeable barriers

INTRODUCTION Pollution generated by industrial activities often causes social, health, and environmental problems, such as the contamination of groundwater by organic and inorganic pollutants. The installation of permeable barriers (PBs) downstream of a contamination plume can remediate affected groundwater more cheaply than the conventional method of excavating the contaminated soil. Furthermore, PBs can passively remediate groundwater and prevent groundwater contamination even in extreme environments (Snape et al. ; Amos & Younger ). Thus, they are viable alternatives to the established pump-and-treat technology. The main component of a PB is a reactive or sorbent material. However, no single material can remove all types of pollutants. Therefore, a material with the ability to remove the contaminants present at the site in question has to be selected when installing any PB system. Other doi: 10.2166/wst.2014.192

factors that inﬂuence the choice of sorbent include the reactivity, hydraulic conductivity, environmental safety, availability, cost, and long-term stability of the sorbent material itself. Several materials are commonly used in the construction of PBs. Zero-valent iron (ZVI) is most frequently used, followed by granular activated carbon. Other materials used as sorbents include microorganisms, natural zeolites, peat, phosphates, limestone and amorphous ferric oxide (Richardson & Nicklow ). Aliphatic hydrocarbons, which are common pollutants at oil-contaminated sites, can be removed by peat (Guerin et al. , ), zeolites (Northcott et al. ; Vignola et al. ), activated carbons (Hornig et al. ) or biologically active barriers (Kao & Wang ; Vesela et al. ). Polyvalent heavy metals (e.g. Cr, As) can be removed by

I. Kozyatnyk et al.

Evaluation of barrier materials for removing pollutants from groundwater

exposure to ZVI. Redox reactions taking place at the iron surface affect the solubility of the heavy metals, which are then sequestered through precipitation or adsorption on iron oxides (Lee & Wilkin ; Mak & Lo ). Metals can also be removed by ion-exchange and adsorption on ﬂy ash (González et al. ; Prasad & Mortimer ), zeolites (Vignola et al. ; Prasad et al. ), mining products (Smyth et al. ; Conca & Wright ) or activated carbon (Lakatos et al. ). In the presented study we evaluated the ability of several PB materials to remove heavy metals and organic pollutants (aliphatic and aromatic hydrocarbons) from groundwater at a 10 ha industrial site in northern Sweden. The soil at the site is also rich in natural organic matter (NOM), particularly fulvic substances, and an adjacent river drains untreated groundwater from the site into the Bothnian Sea. The study is part of a broader project to develop economical and environmentally sustainable methods to prevent the migration of hazardous compounds from both active and abandoned industrial sites.

METHODS Water with high levels of dissolved metals and organic pollutants was extracted for the sorption experiments from two groundwater wells at a 10 ha industrial site in northern Sweden. Water from both wells was pumped into and mixed in a 1 m3 tank, which was flushed with an argon–nitrogen gas mixture to prevent oxidation of metal ions. On the first day of each experiment the gas outlet was situated above the water-line; however, for the remainder of the experiment the gas outlet was submerged. The chemical properties of the composite water are summarized in Table 1 and compared to the water quality recommendation of the European council directive 98/83/EC (European Council ), as well as to the Swedish Geological Survey’s regulations on groundwater status classification and environmental quality standards for groundwater, SGU-FS 2008:2 ().

Experimental design Five sorbent materials were selected based on previous studies on NOM adsorption processes (Dries et al. ; Viipsi et al. ; Liu & Lo ) and results of ﬁeld experiments on the removal of metal and organic contaminants (Richardson & Nicklow ). The materials used in this study are summarized in Table 2.

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The surface area and pore size distribution of each tested material in the dry state were determined by the conventional Brunauer–Emmett–Teller (BET, Brunauer et al. ) and Barrett–Joyner–Halenda (BJH, Barrett et al. ) methods. Prior to analysis, ∼0.5 g of material was transferred to a 9.5 mm (inner diameter) sample tube and degassed at 120 C for 2 h in a Micromeritics SmartPrep unit. A TriStar 3000 automated nitrogen sorption/desorption instrument (Micromeritics, Norcross, GA, USA) was used for 12-point determinations of the BET specific surface area between 0.05 and 0.28 relative pressure, while the BJH pore size distribution was determined by measurements at 77 points (Wikberg et al. ). Six glass columns (internal diameter 19 cm) were packed with sorbent material. Each column was packed with a 5 cm layer of sand at the bottom, followed by a 40 cm thick layer of a 1:1 mixture of a single sorbent material and sand, and finally a 5 cm layer of sand on top, resulting in an overall bed height of 50 cm. Columns containing mixtures of peat, fly ash and sand (1:1:1) were prepared in a similar manner, with the mixture packed between two layers of sand. The end fittings of the columns contained grooves to distribute the flow evenly and were equipped with a fine metal mesh to keep the sorbent material in place. The polluted water was pumped through the columns materials (from bottom up) and the pollutant reduction was evaluated. The experiments were conducted using three parallel columns and, thus, two sets of experiments were required, see Figure 1. Samples of the percolate were taken out in a sequence representing different liquid to sorbent volume ratios (L/S), i.e. 0, 0.5, 2 and 10. Pump flow through the columns was set at L/S ¼ 1 per 24 hours. Water coming out from the columns was collected in glass bottles which had been flushed with 2 mL of concentrated HCl in an attempt to get a pH of the liquor below 3.5. The water samples were stored in a refrigerator room with a temperature of around þ5 C (±2 C). At the end of each experiment, the remaining column content was carefully extracted and four samples were taken (at 10, 20, 30 and 40 cm bed height), for further analysis. W

Sample analysis All collected samples were sent to a commercial laboratory for analysis (Euroﬁns Environment Sweden, Lidköping, Sweden). The substances measured in liquid samples

Table 1

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Evaluation of barrier materials for removing pollutants from groundwater

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Chemical properties of groundwater before and after treatment with six different PB sorbents. For comparison, water quality standards set out by European Council Directive 98/83/EC (European Council 1998) and Swedish Geological Survey SGU-FS 2008:2 (2008) are included, where available Concentrations in ﬁltrate water (ﬁnal sampling)

Quality standards

Raw water

Water in tank before treatment

Fly ash

Peat

Sand

Fly ash/ Peat

Iron

Lignin

Aliphatic (C10C12), mg L 1

2.5

0.19

<0.020

0.045

<0.020

0.056

0.091

Aliphatic (C12C16), mg L 1

3.0

0.30

<0.020

0.092

Aliphatic (C16C35), mg L 1

0.81

<0.050

Fe, mg L 1

129

2.6

<0.010

0.16

<0.020

8.2

0.2

As, mg L

0.35

0.0033

0.0018

0.011

0.004

0.0079

0.0022

<0.0040

0.010

Ba, mg L 1

3.4

0.57

–

Cd, mg L 1

0.011

<0.010

<0.0001

0.00029

0.00012

<0.0001

0.00014

0.005

Co, mg L 1

0.11

<0.0010

0.0012

0.032

0.0057

0.019

0.008

0.042

Cr, mg L 1

0.17

0.0022

<0.0001

0.0024

0.0026

<0.0001

0.015

0.15

0.050

2.9

<0.0030

0.0056

0.0073

0.020

0.055

0.084

0.026

0.002

Hg, mg L 1

0.060

<0.00004

<0.00001

0.000007

0.000011

0.000025

<0.00001

0.001

Mn, mg L 1

3.8

2.5

<0.0010

4.3

3.1

0.01

0.23

6.1

0.050

Ni, mg L 1

0.25

0.0026

0.0032

0.019

0.0051

0.058

0.012

0.021

0.020

3.1

<0.0030

<0.0005

0.0012

<0.0005

0.00052

0.016

0.00086

0.010

–

<0.0020

<0.0005

0.015

0.011

0.0011

0.0039

–

6.5

<0.0050

0.0066

0.012

<0.0050

0.0072

0.036

–

100a

TOC, mg L 1

144

142

229

167

727

DOC, mg L 1

164

152

704

7.5

7.2

12.4

7.6

8.0

10.6

2.2

1.7

6.5– 9.5

Conductivity, mS cm 1

1.6

6.9

1.7

1.8

4.2

2.5

Alkalinity, mmol L 1

9.6

Property

Cu, mg L

Pb, mg L

V, mg L

Zn, mg L

98/83/ EC

SGU-FS 2008:2

0.5a

a Sand and gravel. TOC: total organic carbon; DOC: dissolved organic carbon.

Table 2

Physical properties of the sorbent materials in the experimental PB columns

Material

Supplier

Density, g cm 3

Surface area, m2 g 1

Pore volume, cm3 g 1

Sand, 3 mm fraction

Finja Betong AB, Hässleholm, Sweden

1.60

–

Peat

Hörneborgsverket, Övik Energy AB, Sweden

0.34

0.98

0.009

Fly ash

Hörneborgsverket, Övik Energy AB, Sweden

0.76

Iron, HCA-150N cast iron powder

Hepure Technologies Inc., USA

7.80

1.3

0.009

Lignin

SEKAB-pilot plant, Örnsköldsvik, Sweden

0.70

–

0.051

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Evaluation of barrier materials for removing pollutants from groundwater

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in 300-mL glass ﬂasks, ﬁlled with 100 mL of FA or HA solution and shaken for 72 hours. Equilibrium adsorption was calculated as follows: q¼

ðC0 C Þ V m

(1)

where С0 and С are the concentrations (mg C L 1) of the FA solution before and after the adsorption experiment, V is the solution volume (L), and m is the sorbent mass (g). Assessment of adsorption characteristics of sorbent samples for adsorption from aqueous solutions of FA was carried out using the Freundlich model: q ¼ KF C 1=n

(2)

where KF and n are the Freundlich constants for a given adsorbate and adsorbent at a particular temperature.

RESULTS AND DISCUSSION

Figure 1

Experimental set-up with three columns and a fraction collector.

included total and dissolved organic carbon (TOC and DOC; following ISO 8245:1999), pH (ISO 10523:2008), alkalinity (ISO 9963-1:1994), conductivity (ISO 7888:1985), mercury (ISO 17852:2008), other metals (ISO 17294-2:2003) and organic pollutants (gas chromatography–mass spectrometry (GC-MS); ISO/EN 17025 accredited). Concentrations of metals and organic pollutants in solid materials were estimated using inductively coupled plasma–atomic emission spectroscopy (ISO 11466:1995) and GC-MS (ISO/EN 17025 accredited), respectively. Determination of equilibrium adsorption constants Reference humic acids (HA) and Nordic fulvic acids (FA) were purchased from Sigma-Aldrich, Switzerland, and the International Humic Substance Society, Norway, respectively. These were used to study the equilibrium adsorption of heavy metals and DOC on the various sorbents. Heavy metal salts were added to reference dissolved organic matter solution in their nitrate forms. Adsorption isotherms were measured at 20 C by varying the sorbent amount from 0.005 to 5.000 g. Concentration of HA/FA was 100 mg L 1. Heavy metal concentration was ca. 100 μg L 1. Measurements were conducted W

The concentrations of organic contaminants were greatly reduced by passing water through columns with the sorbent material (Table 1). For metals the effect of the treatment is less unequivocal. Especially, fly ash, peat and lignin can act as sources of several of the metals studied. In any case the levels in the treated groundwater were close to, or below, the limits stipulated in drinking water regulations. Colour, TOC and DOC The water used in the experiments had a fairly high concentration of total organic matter (ca. 140 mg L1 TOC). Fulvic substances probably constituted most of the organic matter and colouring agents in the water, since no precipitation was observed after decreasing the pH of the water to 1–2. Fly ash was most effective for colour removal, but sand, iron and peat/fly ash also removed some colour from the water. A substantial increase in the colour of the filtrate was observed during washings of the peat and lignin columns, probably due to elution of water-soluble substances in the material. The colour of the filtrate decreased towards the end of the experiments, suggesting that these soluble components were not replenished by breakdown of the sorbent. These observations correspond well with measurements of TOC and DOC (defined here are as the fraction of TOC that passes through a 0.45 μm filter). DOC constituted more than 90% of the TOC in filtrates from all of the

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sorbent materials (Table 1), except peat (72%) and ﬂy ash (58%). Overall, only ﬂy ash and iron were capable of reducing TOC and DOC concentrations below initial concentrations. Filtration through lignin in particular, but also peat, caused an increase in the organic matter content of the water. Although both TOC and DOC decreased over time, the levels remained substantially elevated at the end of the experiment. Therefore, sorbent material based on peat or lignin should be thoroughly rinsed prior to use.

Alkalinity, conductivity and pH Filtration through fly ash and fly ash/peat increased the pH of the water, while filtration through iron and lignin decreased it (Table 1). These observations are consistent with expectations since fly ash is known to raise the pH of water solutions (due to its high CaO content (Ricou et al. ), while lignin and iron respectively reduce pH via the leaching of organic acids and oxidation of Fe0 to ferrous and ferric ions followed by precipitation of ferric oxide-hydroxides. Filtration through peat and sand did not affect the pH of the water. Conductivity is a measure of the concentration and charge of ions present in a liquid. Thus, it can indicate changes in pH (OH and Hþ ion concentrations) in water or overall cation/anion concentrations. In experiments with sorbent materials containing fly ash, the increase in pH and elution of metal cations from the fly ash caused an increase in conductivity at the start of the experiment. Conductivity in the filtrate from columns containing iron and lignin also increased over time, most likely due to changes in pH. The filtrates had lower alkalinity than the untreated water at the end of the experiments with all sorbent materials except one, and zero alkalinity in experiments with three sorbent materials: iron, lignin and fly ash/peat. The exception was fly ash, which yielded filtrates with elevated alkalinity throughout the experiment, although it decreased somewhat towards the end.

Reduction in pollutant concentrations The greatest reduction of pollutant contents occurred during the initial equilibration period. Substantial precipitation of NOM was observed in the water storage tank, despite efforts to prevent aging of the water during storage by creating an inert headspace. Large proportions of the organic pollutants and metals co-precipitated with the NOM. More than 90% of the

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pollutants were removed in this process, with the exception of barium (83%), manganese (34%) and cadmium (9%). The removal of dissolved and suspended pollutants from the equilibrated water sample is described in the following sections. Metals Iron and fly ash materials displayed the ability to effectively remove arsenic from water, in accordance with expectations due to the ion-exchange capacity of fly ash and the known ability of arsenic to react with organic substances, forming As–metal–organic complexes (Lin et al. ). Thus, arsenic can be removed through sorption of NOM onto fly ash. ZVI removes arsenate and arsenite rapidly, by spontaneous adsorption and co-precipitation with Fe(II) and Fe(III) oxides and hydroxides, which are formed during the oxidation of ZVI (Ahn et al. ; Bang et al. ). These sorbent materials are also effective for manganese and cobalt removal. Wilopo et al. () showed that manganese can be immobilized through adsorption onto ZVI and by precipitation as carbonates. Lead, chromium and nickel were removed most effectively by sand and sorbents containing fly ash. Sorbents containing fly ash also efficiently removed iron ions, due to precipitation of iron hydroxides under the alkaline conditions created by the fly ash. Not surprisingly, elution of iron was observed after filtration through iron sorbent. An increase of organic matter content after treatment with peat and lignin also led to an increased concentration of iron ions. The reductive removal of Cr(VI) by Fe0 occurs via a surface-mediated electron transfer process, in which Cr(VI) is reduced to Cr(III) by the oxidation of Fe(0) to Fe(II) and Fe(III) and subsequently precipitates out as sparingly soluble Cr, or mixed Fe-Cr, oxyhydroxides (Eary & Rai ; Blowes et al. ). Organic pollutants The most abundant organic pollutants were aliphatic hydrocarbons, followed by aromatic hydrocarbons, polycyclic aromatic hydrocarbons and polychlorinated biphenyls. All sorbents reduced levels of the organic contaminants to below the limits of detection, except for small aliphatic hydrocarbons with fewer than eight carbon atoms (24% breakthrough for sand, 29% for iron and 48% for lignin). Thus, fly ash, peat and fly ash/peat were the most effective sorbents for organic pollutant removal.

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Characterization of adsorbent materials Metal concentrations in unused and used ﬁlter material from different heights in the columns were also determined.

Figure 2

Table 3

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There was little difference in metal concentrations at different heights, so the results for each column were combined and the mean concentrations were used in comparisons of the sorbent materials (Figure 2).

Metal concentrations in ﬁltration materials expressed as changes relative to initial concentrations in the water (%).

Adsorption properties of materials

Fly ash K

Property

Peat n

Sand n

Fe As Ba Cd Cr Cu Hg Pb Zn DOC

0 1.9 –a 0.10 0 11 1.7 4.6 0 0.16

0.44 1.5 –a 0.37 0.18 3.4 1.3 0.74 0.29 0.69

– 0 22 0.72 0 0 –a 3.7 0.17 –a

– 0.03 2.0 0.50 0.13 0.17 –a 0.77 0.67 –a

0.01 0 –b –b –b –b 0 –b –b –b

0.31 0.22 –b –b –b –b 0.13 –b –b –b

– 0.37 0.03 0.09 0.02 1.2 1.3 0.05 0 7.1

–a 0.50 0.15 0.41 0.31 0.62 0.33 0.53 0.18 0.14

Fe As Ba Cd Cr Cu Hg Pb Zn DOC

0 6.9 –a 7.4 –a 8.1 0.85 35 0 0.16

0.52 1.5 –a 2.6 –a 1.6 0.56 5.8 0.44 0.65

–a –a 5.7 0 –a 0.10 –a 0.96 0.09 –a

–a –a 1.6 0.01 –a 0.91 –a 1.4 0.61 –a

0 6.6 –b 0 –b –b 0 0 –b –b

0.21 0.02 –b 0.14 –b –b 0.13 0.18 –b –b

–a 2.8 0.06 0.37 33 4.6 3.6 7.6 0.22 0.17

–a 0.52 0.28 0.50 1.7 0.77 0.34 1.4 0.46 0.29

Elution. No adsorption.

Iron

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Evaluation of barrier materials for removing pollutants from groundwater

In cases where the relative metal concentration exceeded 100%, the metal had been accumulated by the sorbent material. A relative metal concentration of less than 100% indicates that metal had been eluted from the sorbent material. Chromium and vanadium were accumulated by fly ash, whereas other metals were eluted from this material. These observations are similar to findings of Doherty et al. (). Peat materials effectively removed barium, nickel and vanadium. The mixture of peat and fly ash performed better than the individual sorbents, in particular for barium. Several metals (e.g. As, Pb, Cu, Cr, Ni, Mn, Zn, V) eluted from the iron sorbent, due to corrosion and leaching of metal impurities. The sorption behaviour of DOC and the most important heavy metals was investigated using all materials, except lignin, under static conditions. Lignin was excluded because it showed less promise in column studies. Most of the obtained isotherms were of the III-type (S-shape), according to BET classification, which means that a poly-molecular layer was formed during adsorption (Sing et al. ; Giles et al. ). It is impossible to describe these using the Langmuir equation that assumes mono-molecular adsorption and the adsorption was therefore fitted to the general Freundlich equation (Equation (2)). The results are presented in Table 3. Although the Freundlich isotherm equation is empirical, one source (Snoeyink ) showed that KF is related to the capacity of adsorbent for the adsorbate and n is related to the strength of adsorption driving force (energy of adsorption). Assessing the n coefficients of the Freundlich equation we can conclude that fly ash was the most effective sorbent for DOC and heavy metals. Iron also had promising removal properties. However, peat was the most effective sorbent for barium and zinc. Sand was a weak sorbent under static conditions, in line with the results of the column studies.

CONCLUSIONS Heavy metals were most effectively removed by sorbents based on fly ash and zero-valent iron, whereas aliphatic compounds were most effectively removed by sorbents based on fly ash and peat. Static sorption studies indicate that fly ash provides the strongest sorption of the tested sorbents. Lignin and peat sorbents leach during filtration, which causes increased TOC concentrations in the groundwater. Therefore, lignin and peat sorbents should be washed prior to use. Aeration seems to be efficient in removing pollutants,

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especially heavy metals. Remediation of contaminated land at disused industrial sites is often a complex task that requires the use of mixed sorbent materials, or two or more sequential barriers, or sorption combined with aeration.

ACKNOWLEDGEMENT Financial support from the Kempe Foundation is gratefully acknowledged.

REFERENCES Ahn, J. S., Chon, C.-M., Moon, H.-S. & Kim, K.-W.  Arsenic removal using steel manufacturing byproducts as permeable reactive materials in mine tailing containment systems. Water Research 37 (10), 2478–2488. Amos, P. W. & Younger, P. L.  Substrate characterisation for a subsurface reactive barrier to treat colliery spoil leachate. Water Research 37 (1), 108–120. Bang, S., Johnson, M. D., Korfiatis, G. P. & Meng, X.  Chemical reactions between arsenic and zero-valent iron in water. Water Research 39 (5), 763–770. Barrett, E. P., Joyner, L. G. & Halenda, P. P.  The determination of pore volume and area distributions in porous substances. I. Computations from nitrogen isotherms. Journal of the American Chemical Society 73 (1), 373–380. Blowes, D. W., Ptacek, C. J., Benner, S. G., McRae, C. W. T., Bennett, T. A. & Puls, R. W.  Treatment of inorganic contaminants using permeable reactive barriers. Journal of Contaminant Hydrology 45 (1–2), 123–137. Brunauer, S., Emmett, P. H. & Teller, E.  Adsorption of gases in multimolecular layers. Journal of the American Chemical Society 60 (2), 309–319. Conca, J. L. & Wright, J.  An Apatite II permeable reactive barrier to remediate groundwater containing Zn, Pb and Cd. Applied Geochemistry 21 (8), 1288–1300. Doherty, R., Phillips, D. H., McGeough, K. L., Walsh, K. P. & Kalin, R. M.  Development of modified flyash as a permeable reactive barrier medium for a former manufactured gas plant site, Northern Ireland. Environmental Geology 50 (1), 37–46. Dries, J., Bastiaens, L., Springael, D., Kuypers, S., Agathos, S. N. & Diels, L.  Effect of humic acids on heavy metal removal by zero-valent iron in batch and continuous flow column systems. Water Research 39 (15), 3531–3540. Eary, L. E. & Rai, D.  Chromate removal from aqueous wastes by reduction with ferrous ion. Environmental Science & Technology 22 (8), 972–977. European Council  Council directive 98/83/EC of 3 November 1998 on the quality of water intended for human consumption. Official Journal of the European Communities L 330, 0032–0054.

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Giles, C. H., Smith, D. & Huitson, A.  A general treatment and classification of the solute adsorption isotherm. I. Theoretical. Journal of Colloid and Interface Science 47 (3), 755–765. González, A., Navia, R. & Moreno, N.  Fly ashes from coal and petroleum coke combustion: current and innovative potential applications. Waste Management & Research 27 (10), 976–987. Guerin, T. F., McGovern, T. & Homer, S.  A funnel and gate system for remediation of dissolved phase petroleum hydrocarbons in groundwater. Land Contamination and Reclamation 9 (2), 209–224. Guerin, T. F., Horner, S., McGovern, T. & Davey, B.  An application of permeable reactive barrier technology to petroleum hydrocarbon contaminated groundwater. Water Research 36 (1), 15–24. Hornig, G., Northcott, K., Snape, I. & Stevens, G.  Assessment of sorbent materials for treatment of hydrocarbon contaminated ground water in cold regions. Cold Regions Science and Technology 53 (1), 83–91. Kao, C. M. & Wang, C. C.  Control of BTEX migration by intrinsic bioremediation at a gasoline spill site. Water Research 34 (13), 3413–3423. Lakatos, J., Brown, S. D. & Snape, C. E.  Coals as sorbents for the removal and reduction of hexavalent chromium from aqueous waste streams. Fuel 81 (5), 691–698. Lee, T. R. & Wilkin, R. T.  Iron hydroxy carbonate formation in zerovalent iron permeable reactive barriers: Characterization and evaluation of phase stability. Journal of Contaminant Hydrology 116 (1–4), 47–57. Lin, H. T., Wang, M. C. & Li, G. C.  Complexation of arsenate with humic substance in water extract of compost. Chemosphere 56 (11), 1105–1112. Liu, T. & Lo, I. M. C.  Influences of humic acid on Cr(VI) removal by zero-valent iron from groundwater with various constituents: Implication for long-term PRB performance. Water, Air, and Soil Pollution 216 (1–4), 473–483. Mak, M. S. H. & Lo, I. M. C.  Influences of redox transformation, metal complexation and aggregation of fulvic acid and humic acid on Cr(VI) and As(V) removal by zerovalent iron. Chemosphere 84 (2), 234–240. Northcott, K. A., Bacus, J., Taya, N., Komatsu, Y., Perera, J. M. & Stevens, G. W.  Synthesis and characterization of hydrophobic zeolite for the treatment of hydrocarbon contaminated ground water. Journal of Hazardous Materials 183 (1–3), 434–440. Prasad, B. & Mortimer, R.  Treatment of acid mine drainage using fly ash zeolite. Water, Air, & Soil Pollution 218 (1), 667–679. Prasad, B., Sangeeta, K. & Tewary, B. K.  Fly ash zeolite as permeable reactive barrier for prevention of groundwater contamination due to coal ash disposal. Asian Journal of Chemistry 24 (3), 1045–1050.

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Richardson, J. P. & Nicklow, J. W.  In situ permeable reactive barriers for groundwater contamination. Soil and Sediment Contamination: An International Journal 11 (2), 241–268. Ricou, P., Lécuyer, I. & Le Cloirec, P.  Removal of Cu2þ, Zn2þ and Pb2þ by adsorption onto fly ash and fly ash/lime mixing. Water Science and Technology 39 (10–11), 239–247. SGU-FS 2008:2  Sveriges geologiska undersöknings föreskrifter om statusklassificering och miljökvalitetsnormer för grundvatten (Swedish Geological Survey regulations on groundwater status classification and environmental quality standards for groundwater). Sveriges geologiska undersöknings författningssamling, Swedish Geological Survey, Uppsala, Sweden. Sing, K. S. W., Everett, D. H., Haul, R. A. W., Moscou, L., Pierotti, R. A., Rouquerol, J. & Siemieniewska, T.  Reporting physisorption data for gas solid systems with special reference to the determination of surface-area and porosity (Recommendations 1984). Pure and Applied Chemistry 57 (4), 603–619. Smyth, D., Blowes, D., Ptacek, C. & Bain, J.  Application of permeable reactive barriers for treating mine drainage and dissolved metals in groundwater. Geotechnical News 22 (1), 39–44. Snape, I., Morris, C. E. & Cole, C. M.  The use of permeable reactive barriers to control contaminant dispersal during site remediation in Antarctica. Cold Regions Science and Technology 32 (2–3), 157–174. Snoeyink, V. L.  Adsorption of organic compounds. In: Water Quality and Treatment, 4th edn (F. W. Pontius, ed.). McGraw-Hill, NewYork, pp. 781–785. Vesela, L., Nemecek, J., Siglova, M. & Kubal, M.  The biofiltration permeable reactive barrier: practical experience from Synthesia. International Biodeterioration and Biodegradation 58 (3–4), 224–230. Vignola, R., Bagatin, R., De Folly D’Auris, A., Flego, C., Nalli, M., Ghisletti, D., Millini, R. & Sisto, R.  Zeolites in a permeable reactive barrier (PRB): One year of field experience in a refinery groundwater – Part 1: The performances. Chemical Engineering Journal 178, 204–209. Viipsi, K., Tõnsuaadu, K. & Peld, M.  Impact of soluble humic substance on Cd2þ sorption on apatite in aqueous solutions. Chemistry and Ecology 26 (Suppl. 2), 77–85. Wikberg, E., Sparrman, T., Viklund, C., Jonsson, T. & Irgum, K.  A 2H nuclear magnetic resonance study of the state of water in neat silica and zwitterionic stationary phases and its influence on the chromatographic retention characteristics in hydrophilic interaction high-performance liquid chromatography. Journal of Chromatography A 1218 (38), 6630–6638. Wilopo, W., Sasaki, K., Hirajima, T. & Yamanaka, T.  Immobilization of arsenic and manganese in contaminated groundwater by permeable reactive barriers using zero valent iron and sheep manure. Materials Transactions 49 (10), 2265–2274.

First received 5 December 2013; accepted in revised form 7 April 2014. Available online 19 April 2014

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Deposit membrane fouling: influence of specific cake layer resistance and tangential shear stresses A. Charfi, J. Harmand, N. Ben Amar, A. Grasmick and M. Heran

ABSTRACT Cake fouling is the leading cause of membrane permeability decrease when filtering mixed liquor suspension containing high suspended solid concentrations. A simple model is proposed to simulate the cake resistance evolution with time by considering a macro-scale fouling linked only to the accumulation of particles on the membrane surface. This accumulation appears as the difference between the flux of deposited particles due to the filtration and the flux of particles detached from the membrane surface due to the tangential shear stresses caused by recirculation flow in the sidestream membrane bioreactor (MBR) or gas sparging close to the membrane surface for submerged MBR configuration. Two determining parameters were then highlighted: the specific cake resistance and the ‘shear parameter’. Based on these parameters it is possible to predict model outputs as cake resistance and permeate flux evolution for short-time filtration periods. Key words

| cake formation, membrane fouling modelling, speciﬁc cake resistance, tangential shear stresses

NOMENCLATURE macc

Speciﬁc mass of accumulated cake (g m 2)

macc,lim

Maximum speciﬁc mass of accumulated cake (g m 2)

min

Speciﬁc mass of particles approaching the membrane surface (g m 2)

mout

Speciﬁc mass of particles removed from the membrane by shear forces (g m 2)

Total suspended solids concentration (gTSS L 1)

Permeate ﬂux (L m 2 h 1)

TMP

Trans-membrane pressure (Pa)

Dynamic viscosity (Pa s)

Clean membrane resistance (m 1)

Cake resistance (m 1)

Initial permeate ﬂux (L m 2 h 1)

Shear parameter (m2 kg 1)

Speciﬁc cake resistance (m kg 1)

Rc,max

Resistance at steady state (m 1)

Jp,end

Permeate ﬂux at steady state (L m 2 h 1)

doi: 10.2166/wst.2014.186

A. Charfi (corresponding author) A. Grasmick M. Heran Institut Européen des Membranes (IEM), Université Montpellier 2, place Eugene Bataillon, CC05, 34095 Montpellier, France E-mail: amine.charfi@ymail.com A. Charfi N. Ben Amar Laboratoire de Modélisation Mathématique et Numérique dans les Sciences de l’Ingénieur, ENIT, B. P 37 Le Belvédère 1002 Tunis, Tunisia and Institut National des Sciences Appliquées et de Technologie, B. P. 676, 1080 Tunis Cedex, Tunisia J. Harmand Laboratoire de Biotechnologie de l’Environnement, INRA, UR0050, Avenue des étangs, 11100 Narbonne, France and MODEMIC Research Group, INRIA, Laboratoire de Mathématiques, Informatique et Statistique pour l’Environnement et l’Agronomie-INRA, 2 Place Viala, 34060 Montpellier, France

INTRODUCTION Many studies have tried to model fouling in membrane processes. Many are based on the four basic fouling mechanisms introduced by Hermia (), namely cake fouling, pore constriction, complete blocking, and intermediate blocking. Despite being established for dead-end filtration at constant pressure, these models have been also used for cross-flow filtration (Ho & Sung ; Charfi et al. ) or have been adapted to cross-flow filtration (Field et al. ) by taking into account the role of shear forces to eliminate particle deposition on the membrane surface (De Bruijn et al. ). More realistic models are also presented when considering simultaneous fouling mechanisms; while Bolton et al. () combined Hermia’s models, Wu et al. () established a model considering simultaneously three fouling mechanisms, namely pore constriction, pore blockage, and cake filtration associated to soluble, colloidal and suspended activated sludge components, respectively. Abdelrasoul et al. () considered the probability of particle deposition on the membrane surface or on another

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particle to predict the mass of particles accumulated on the membrane surface, on membrane pores and inside pores. Other authors considered the resistance in series model based on Darcy’s law in which the resistance term is the sum of resistances associated with the membrane, the internal clogging and the cake deposit (Song ; Ho & Zydney ; Kuberkar & Davis ; Katsoufidou et al. ). A concentration polarization resistance term was also proposed (Choi et al. ). An adsorption resistance term could be also inserted when affinity is important between effluent components and membrane material (Bhattacharjee & Bhattacharya ; Vela et al. ). Li & Wang () considered the uneven shear stress on the membrane surface leading to uneven cake deposit, and modelled fouling based on biomass attachment and detachment. Many studies are based on the critical flux approach proposed by Field et al. () who assume that avoiding membrane fouling would be possible if working under a so-called critical flux for which no trans-membrane pressure (TMP) increase is detected. However, even in sub-critical flux operation, a two-step TMP increase would be observed. Ognier et al. () proposed a model describing TMP increase assuming that a progressive pore blocking would lead to a local flux increase and consequently to membrane fouling increase. In face of the difficulty of modelling so complex an interaction between suspension components and membrane structure and its associated dynamic layers formed by accumulation of compounds close to the membrane surface, it was decided to propose a simple model of membrane fouling by defining parameters in link with tangential shear stresses. Such a simple model is proposed in this work. Considering mainly the case of filtration of activated sludge suspensions, the main fouling mechanism in this model is supposed to be cake formation due to compound accumulation onto the membrane surface during filtration (Charfi et al. ). This kind of model is typically intended to complete those developed for bioprocess control design such as the modified two-step anaerobic digestion model (AM2b) proposed by Benyahia et al. () for anaerobic membrane bioreactors. Such a model is proposed hereafter.

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TMP of 0.5 bar and with a recirculation velocity of 1.7 m s 1. No cleaning operations have been applied during experiments. Nevertheless membranes were cleaned chemically after each experiment. The influents were digestate effluents extracted from two different anaerobic digesters treating agricultural wastes. It was either directly filtered (raw digestate) or pre-treated by centrifugation (Dig A and Dig B) for 15 min under a relative centrifugation force of 2,500 g.

THE CAKE MODEL Model development In this section, the assumed hypotheses are presented and mass balances are established to develop the basic relations of the model. Hypothesis The size of particles present in the suspension is assumed to be larger than the membrane pores. Given that particles are totally retained by the membrane barrier, it is then assumed that membrane fouling is only due to particular accumulation on membrane surface. The only fouling mechanism considered is cake formation. Considering a constant TMP condition, the membrane fouling would be described by permeate flux decrease according to the hydraulic resistance increase with time (Darcy’s law). Mass balance The establishment of mass balance on the membrane surface (Equation (1)) should allow the determination of the accumulated cake mass macc per membrane surface unit at time t and its corresponding influence on permeate flux decline. During a short period of filtration, dt, the specific mass dmacc of particles accumulated can be estimated as the difference between the specific mass dmin of particles carried on to the membrane surface by permeation forces during filtration and the mass dmout of particles dragged away from the membrane surface by shear forces due to tangential flow:

MATERIAL AND METHODS dmacc =dt ¼ dmin =dt dmout =dt A cross-ﬂow ﬁltration set-up (Membralox) was used. Membranes used were mono-tubular ceramic membranes with an area of 55 cm2 and pore sizes of 0.8, 0.05, and 0.02 μm. Temperature was maintained at 37 C by thermostatic bath. Experiments of 80 min were conducted under a constant W

(1)

At time t, the specific mass brought by filtration is considered as proportional to the concentration X of suspended solids (supposed as a constant value) and to the flux of permeation, as shown in (Equation (2)), where Jp is

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the speciﬁc permeate ﬂux at time t: dmin ¼ X:Jp dt

(2)

The detached specific mass of particle within a short period of time due to parietal shear stresses is assumed to be proportional to (i) the specific mass macc of particle accumulated at time t (supposed proportional to the instantaneous cake thickness) and (ii) the specific flux of particles brought to membrane surface by permeation dmin/dt (Equation (3)). This relation is based on the following assumptions: (i) when macc is low, the detachment is weak, and (ii) when macc has reached its maximal value (corresponding to the laminar boundary layer thickness), the detached flux is equal to the particle flux coming onto the membrane surface due to filtration. In this relation, β appears as a coefficient representing the capacity of parietal shear stresses to detach deposited particles:

dmout dmin macc dmin ¼β ¼ : :macc dt dt macc, lim dt

(3)

In Equation (3), macc,lim is assumed to be an inverse function of local shear stresses, expressed through the criterion β (with β ¼ 1/macc,lim). The variation of the accumulated particle mass per membrane surface unit is then expressed by Equation (4): dmacc ¼ XJp ð1 βmacc Þ dt

(4)

To describe the permeate ﬂux Jp, a resistance in series model based on Darcy’s law is used (Equation (5)): Jp ¼

TMP μ ðR0 þ Rc Þ

(5)

where μ is the dynamic viscosity of permeate, R0 is the clean membrane resistance, and Rc is the cake resistance. Rc and R0 are expressed in Equations (6) and (7), respectively: Rc ¼ αmacc

(6)

TMP μJ0

(7)

R0 ¼

with α and J0 are the speciﬁc resistance of the cake deposit and the initial permeate ﬂux obtained with the clean membrane, respectively.

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Equations (4) and (8) show the basic relation of the cake model finally obtained. Two parameters are then identified as α and β. Their knowledge allow the simulation of the evolution of the specific cake mass in the studied system, the value of the cake resistance and the evolution of the permeate flux with time: 8 dmacc > > ¼ XJp ð1 βmacc Þ > > dt < TMP Jp (t) ¼ > > TMP > > : þ μαmacc (t) J0

(8)

Numerical simulation Parameter values were determined using the experimental data by the least square method programmed with MATLAB R2008b. Qualitative behaviour A preliminary study of the qualitative model behaviour was conducted. Numerical simulations (Figure 1) confirm a cake mass increasing with time until reaching a stationary value corresponding to an equilibrium between accumulation of particles due to filtration and their detachment due to parietal shear stresses; this phenomenon induces a permeate flux decreasing with time until reaching a stationary value. An increasing α induces a decrease of permeate flux and an increase of cake resistance. An increasing β induces an increase of permeate flux which reaches more quickly its stationary value and a decrease of the deposit mass and cake resistance. Model predictions of the stationary conditions Figure 2 gives an illustration of the influence of α and β parameters on two practical key variables Rc,max and Jp,end which represent the stationary values obtained at the equilibrium between deposition and detachment of particles on the membrane surface. Figure 2 clearly indicates the following points: (i) when α is low (<1013 m kg 1), the cake resistance appears low and it is not necessary to develop high shear stresses to maintain a high level of permeate flux, and (ii) in contrast when α is high (>1014 m kg 1), the influence of β is evident but its efficiency could be rapidly limited for very high α values except by combining parietal shear stresses and other ways of

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Deposit membrane fouling

Water Science & Technology

Figure 1

Model simulations of macc and Jp evolutions with time (α1 ¼ 1013, α2 ¼ 1014 m kg 1, and β1 ¼ 10, β2 ¼ 400 m2 kg 1).

Figure 2

Obtained cake resistance and permeate ﬂux when varying the parameters α and β.

surface cleaning (relaxation period or back-flushing for example). To study their effect on the model output, a sensitivity study of the parameters specific cake resistance α (m kg 1) and shear parameter β (m2 kg 1), was conducted. The variations of parameters α and β lead to a considerable variation of the cake resistance and the permeate flux at stationary phase noted as Rc,max and Jp,end. These outputs are sensitive to parameters α and β, which should be well optimized to control the system. Parameter identification and model validation Parameter identification The two model parameters, specific cake resistance α (m kg 1) and shear parameter β (m2 kg 1), are hard to accurately determine experimentally.

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Two approaches have been developed. The former considered the experimental measurement of α by a specific dead-end mode filtration and β was obtained by adjustment of simulated stationary Jp value with experimental data. The second considered a global simulation; the optimal parameter values were obtained by least squares method when comparing experimental data and simulation of Jp evolution versus time (Equation (8)). Figure 3(a) represents the Jp evolutions obtained: (i) experimentally when filtering an effluent from an anaerobic digester with a microfiltration (MF) membrane (0.8 μm), (ii) by simulation (S1) carried out by using α measured by deadend filtration, and (iii) by simulation (S2) when α and β were directly calculated by global simulation. It is clearly obvious that the simulation taking into account the experimental value of α measured in dead-end filtration (1015 m kg 1) did not allow the representation of the evolution of Jp even when choosing a very high value

Figure 3

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Deposit membrane fouling

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Simulation and experimental data: (a) centrifuged digestate (Dig A) ﬁltered using membrane (0.8 μm); (b), (c) and (d) raw digestate ﬁltered with membrane pore sizes 0.05, 0.02, and 0.8 μm, respectively.

of β (more than 104 m2 kg 1). This point confirms the fact that the deposit obtained in tangential mode has not the same characteristics, notably in terms of specific hydraulic resistance (Gasmi ). Despite the fact that shear forces are supposed to drag out the biggest particles and form denser cake deposit, the specific cake resistance of such deposit obtained in tangential mode appears lower than when working in dead-end mode that quickly generates an important accumulation of matter presenting a very high specific hydraulic resistance. In contrast, the global approach of simulation made possible the determination of α and β values with a qualitatively good fitting to the experimental data. Evaluation of the model To evaluate the model, three sets of experimental data have been considered; these data were obtained while filtering under the same operating conditions (same shear stress

conditions) the same anaerobic digestate (raw digestate) by means of three different membrane pore sizes (0.8, 0.05 and 0.02 μm). When filtering, an important decrease of permeate flux was observed (Figures 3(a)–3(d)), due to a quick build-up of deposit on the membrane surface in accordance with the high concentration of suspended solids (≈40 g L 1) in the suspension, as indicated by Ho & Sung (). According to the filtering conditions (same shear stress and same suspension) the formed cake could have similar characteristics (even if the smallest cut-off should retain the smallest compounds). When fitting the model with data obtained using the 0.05 μm membrane (Figure 3(b)), the parameter values found were α ¼ 5.84 × 1013 m kg 1 and β ¼ 237 m2 kg 1. Using these values for the two other experiments (0.02 μm in Figure 3(c) and 0.8 μm in Figure 3(d)), the obtained simulated curves give a qualitatively good representation of Jp evolutions versus time; however, a space between model and experimental data can be observed at the end of

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Table 1

Deposit membrane fouling

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Cake model parameter optimized values

Efﬂuent

Membrane pore size (μm)

gTSS L 1

Dig A

0.8

4 4

Dig A

0.02

Least squares

0.13 × 1014

258.25

4.88 × 10 10

0.97

263.42

5.06 × 10 10

0.85

1.83 × 10

Dig B

0.8

0.88 × 10

265.44

6.05 × 10

0.87

Dig B

0.02

3.37 × 1014

281.12

2.49 × 10 10

0.87

experiment (coefficient of determination R 2 of almost 80%). Nevertheless, as this model should allow the development of a tool optimizing the washing procedure of the membrane during operation, including notably backwashing, it is considered that the model is able to give a good illustration of the cake formation and a sufficient accuracy of its specific resistance during the first 10–20 min of each filtration cycle. Table 1 gives other examples of values of α and β obtained with the same two liquid phases after centrifugation (Dig A and Dig B); the new obtained liquid phases contained lower suspended solids concentration (4 g L 1 instead of 40 g L 1). Because the filtrations were carried out in the same shear stress conditions, β values obtained by simulation are very close for the four experiments. The new specific cake resistances appeared only slightly higher, proving the determining role of the smallest particles remaining in the centrifuged suspension on cake resistance. Moreover α obtained for the ultrafiltration membrane (0.02 μm) appears notably higher than the value obtained with MF membrane (0.8 μm). In fact, the digestate centrifugation kept only small particles in liquid phase; then the membrane cut-off played an important role for small particle selectivity, deposit structuring, and dominant membrane fouling mechanisms. It can be noticed that the values obtained for α are satisfying regarding of the values mentioned in the literature (Lin et al. ). Nevertheless, as shown in Figure 3 and described in the literature, the hydraulic resistance due to compounds’ accumulation on the membrane surface continuously increases with time due to fouling deterioration including biofilm development, pore blocking and adsorption, which should be incorporated to complete this model.

CONCLUSION A simple model is proposed to describe tangential ﬁltration of anaerobic digestate suspension where cake layer deposit is the dominant fouling mechanism. The development of

the proposed model equations led to two determining parameters: one is represented by the specific resistance α of the cake deposit, while the second one, β, represents the influence of parietal shear stresses on the limitation of the cake development on the membrane surface. Model simulations confirm the determining role of α and β. Comparisons between experimental data and simulations confirm the value of such a simple model to determine the specific cake resistance in a cross-flow filtration mode, for which it is hard to quantify experimentally α and β due to the difficulty of observing in-line the cake layer structuring and to quantify the shear stress intensity close to the membrane surface.

ACKNOWLEDGEMENTS The authors would like to acknowledge the ﬁnancial support of the French National Research Agency (ANR–Fisscarden), notably the TREASURE Euromed research network and the competitiveness cluster TRIMATEC, as well as the PHC UTIQUE project no. 13G1120.

REFERENCES Abdelrasoul, A., Doan, H. & Lohi, A.  A mechanistic model for ultrafiltration membrane fouling by latex. Journal of Membrane Science 433, 88–99. Benyahia, B., Sari, T., Cherki, B. & Harmand, J.  Anaerobic membrane bioreactor modeling in the presence of soluble microbial products (SMP) – the anaerobic model AM2b. Chemical Engineering Journal 228, 2011–2022. Bhattacharjee, S. & Bhattacharya, P. K.  Flux decline behaviour with low molecular weight solutes during ultrafiltration in an unstirred batch cell. Journal of Membrane Science 72, 149–161. Bolton, G. R., Boesch, A. W. & Lazzara, M. J.  The effects of flow rate on membrane capacity: development and application of adsorptive membrane fouling models. Journal of Membrane Science 279, 625–634.

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Deposit membrane fouling

Charfi, A., Ben Amar, N. & Harmand, J.  Analysis of fouling mechanisms in anaerobic membrane bioreactors. Water Research 46, 2637–2650. Choi, S. W., Yoon, J. Y., Haam, S., Jung, J. K., Kim, J. H. & Kim, W. S.  Modeling of the permeate flux during microfiltration of BSA-adsorbed microspheres in a stirred cell. Journal of Colloid and Interface Science 228, 270–278. De Bruijn, J. P. F., Salazar, F. & Borquez, R.  Membrane blocking in ultrafiltration: a new approach to fouling. Food and Bioproducts Processing 83, 211–219. Field, R. W., Wu, D., Howell, J. A. & Gupta, B. B.  Critical flux concept for microfiltration fouling. Journal of Membrane Science 100, 259–272. Gasmi, A.  Intérêt d’un bioréacteur à membranes immergées pour le traitement de la pollution azotée dans une usée carencée en matière organique (Interest of submerged membrane bioreactor to treat wastewater containing ammonium salt). PhD thesis, Université Montpellier2, Montpellier, France. Hermia, J.  Constant pressure blocking filtration laws – application to power law non-Newtonian fluids. Transactions IChemE 60, 183–187. Ho, C. C. & Zydney, A. L.  A combined pore blockage and cake filtration model for protein fouling during microfiltration. Journal of Colloid and Interface Science 232, 389–399. Ho, J. & Sung, S.  Effects of solid concentrations and crossflow hydrodynamics on microfiltration of anaerobic sludge. Journal of Membrane Science 345, 142–147. Katsoufidou, K., Yiantsios, S. G. & Karabelas, A. J.  A study of ultrafiltration membrane fouling by humic acids and flux

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recovery by backwashing: experiments and modeling. Journal of Membrane Science 266, 40–50. Kuberkar, V. T. & Davis, R. H.  Modeling of fouling reduction by secondary membranes. Journal of Membrane Science 168, 243–258. Li, X. Y. & Wang, X. M.  Modelling of membrane fouling in a submerged membrane bioreactor. Journal of Membrane Science 278, 151–161. Lin, H., Liao, B. Q., Chen, J., Gao, W., Wang, L., Wang, F. & Lu, X.  New insights into membrane fouling in a submerged anaerobic membrane bioreactor based on characterization of cake sludge and bulk sludge. Bioresource Technology 102, 2373–2379. Ognier, S., Wisniewski, C. & Grasmick, A.  Membrane bioreactor fouling in sub-critical filtration conditions: a local critical flux concept. Journal of Membrane Science 229, 171–177. Song, L.  Flux decline in crossflow microfiltration and ultrafiltration: mechanisms and modeling of membrane fouling. Journal of Membrane Science 139, 183–200. Vela, M. C. V., Blanco, S. A., Garcia, J. L., Gozalvez-Zafrilla, J. M. & Rodriguez, E. B.  Modelling of flux decline in crossflow ultrafiltration of macromolecules: comparison between predicted and experimental results. Desalination 204, 328–334. Wu, J., He, C., Jiang, X. & Zhang, M.  Modeling of the submerged membrane bioreactor fouling by the combined pore constriction, pore blockage and cake formation mechanisms. Desalination 279, 127–134.

First received 17 September 2013; accepted in revised form 7 April 2014. Available online 22 April 2014

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Application of three tailing-based composites in treating comprehensive electroplating wastewater Hongbo Liu, Mengling Zhu and Saisai Gao

ABSTRACT Heavy metals and chemical oxygen demand (COD) are major challenging pollutants for most electroplating wastewater treatment plants. A novel composite material, prepared with a mixture of calcium and sodium compounds and tailings, was simply mixed by ratios and used to treat a comprehensive electroplating wastewater with inﬂuent COD, total copper (T-Cu), and total nickel (T-Ni) respectively as 690, 4.01, and 20.60 mg/L on average. Operational parameters, i.e. the contact

Hongbo Liu (corresponding author) Mengling Zhu Saisai Gao School of Environment and Architecture, University of Shanghai for Science and Technology, 200093 Shanghai, China E-mail: Liuhb@usst.edu.cn

time, pH, mass ratio of calcium and sodium compounds and tailings, were optimized as 30 min, 10.0, and 4:2:1. Removal rates for COD, T-Cu, and T-Ni could reach 71.8, 90.5, and 98.1%, respectively. No signiﬁcant effect of initial concentrations on removal of T-Cu and T-Ni was observed for the composite material. The adsorption of Cu(II) and Ni(II) on the material ﬁtted Langmuir and Freundlich isotherms respectively. Weight of waste sludge from the calcium/sodium–tailing system after reaction was 10% less than that from the calcium–tailing system. The tailing-based composite is costeffective in combating comprehensive electroplating pollution, which shows a possibility of applying the tailings in treating electroplating wastewater. Key words

| copper, electroplating wastewater, heavy metals, nickel, tailing-based composite

INTRODUCTION Heavy metals and many organic substances discharged by industrial activities could be taken and accumulated in tissues of living organisms through the food chain (Sciban et al. ). Toxic metals such as copper, zinc, nickel, and chromium ﬁnd their ways to water bodies through industrial wastes, among which electroplating wastewater is of intensive concern (Kurniawan et al. ). Copper in water comes from various sources such as metal cleaning and plating baths, pulp and paper board mills, and fertilizer industry; it is harmful to human health and aquatic life (Congeevaram et al. ). The highest acceptable concentration of total copper (T-Cu) in drinking water is 1.3 and 2.0 mg/L respectively according to the US Environmental Protection Agency and the World Health Organization in order to avoid excessive intake that could cause chronic copper poisoning, hemochromatosis and gastrointestinal catarrh (Gupta ; Anirudhan & Suchithra ). Nickel in water comes from plating plants and steel factories, and the production of nickel batteries and alloys (Ajmal et al. ). Excessive intake of nickel could cause cancer in lung, nose and bone (Yavuz et al. ). Exposure doi: 10.2166/wst.2014.191

to nickel, such as coins and costume jewelry could also cause dermatitis (Kadirvelu et al. ). According to the Indian Standards Institution, the permissible total nickel (T-Ni) in industrial effluent is 3.0 mg/L (Congeevaram et al. ). Heavy metals and chemical oxygen demand (COD) can be removed from electroplating wastewater by methods such as precipitation, ion exchange, electrochemical reduction, evaporation, and reverse osmosis (Ye et al. ). However, the high cost of these methods is still a common complaint for project owners, especially for those in developing counties (Sousa et al. ). Removing pollutants with natural particulate matter, even with waste materials, is currently believed to be a promising solution to reduce cost (Kurniawan et al. ). Efforts have been orientated to find efficient, cheap and easy-to-operate processes using kaolinite, activated slag, Chlorella sorokiniana, baker’s yeast, and kyanite (Gupta ; Ajmal et al. ; Padmavathy et al. ; Yavuz et al. ; Akhtar et al. ). Removal of nickel and copper by marine algal biomass, grape stalk wastes, anaerobic granular

H. Liu et al.

Application of three tailing-based composites in treating comprehensive electroplating wastewater

biomass, and clay sorbent has also been reported (Vengris et al. ; Sheng et al. ; Villaescusa et al. ; Hawari & Mulligan ). Tailings are wastes that cause environmental problems during the metallurgy and mining processes. It is possible to use tailings that contain useful elements such as manganese, iron, and aluminum to treat organics and heavy metals. In this study a novel composite material was prepared by proportionally mixing calcium/sodium compounds and tailings. The feasibility of using the material in electroplating wastewater treatment was studied through investigating performances of different calcium/sodium-tailing systems and economic assessment. Inﬂuences of operational parameters, i.e. the contact time, pH, mass ratio of calcium and sodium compounds and tailings, on performances were studied. Isothermal adsorption experiments were carried out to reveal pollutant removal mechanisms of the material.

MATERIALS AND METHODS Materials Qualitative ﬁlter paper, cationic polyacrylamide, calcium oxide and sodium hydroxide were purchased from Sinopharm Chemical Reagent Co., Ltd (Shanghai, China). The manganese tailing and copper–iron (Cu/Fe) tailing were obtained from a mineral site in Guangxi China. Stock T-Ni and T-Cu solution (1,000 mg/L) was purchased from Shanghai Institute of Measurement and Testing Technology (Shanghai, China). T-Ni and T-Cu solutions used in this study were diluted from the stock solution with de-ionized water. To avoid interferences of other metal ions, all the glassware such as Erlenmeyer ﬂasks, beakers and measuring cylinder were washed three times by de-ionized water before use. Wastewater The wastewater was collected daily from the acid–nickel– copper line of an electroplating factory located in Jiangsu China. Characteristics of the wastewater are listed in Table 1, and pH was in the range of 1–2. Analytical methods Inductively coupled plasma mass emission spectroscopy (NexION 300) was used to characterize heavy metal components in electroplating wastewater qualitatively. T-Cu

Table 1

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Characteristics of the wastewater

Pollutants

Concentration (mg/L)

T-Ni

19.98 ± 0.45

T-Cu

3.91 ± 0.84

COD NH3-N Total phosphorus

693 ± 72 58.84 ± 1.67 345.09 ± 16.32

and T-Ni were determined by the atomic absorption spectrophotometer (TAS-990) quantitatively. COD was determined by the HACH method (Cat. 2125925) and pH was determined by Mettler Toledo FE20. Ammonium (NH3-N), total phosphorus (TP) and sludge (suspended solids) were determined by standard methods (APHA ). Description of experimental procedures Performances of two sole tailing systems Performances of two kinds of tailings, i.e. the Cu/Fe tailing and the Mn tailing were investigated with parallel tests. For each test a different amount of tailing was mixed with 100 mL wastewater in a 250 mL Erlenmeyer beaker and shaken constantly at 150 rpm for 6 hours. One milligram of cationic polyacrylamide was then dosed into the suspension and slowly shaken for 5 min; the supernatant was drawn out for analysis after precipitation for 30 min. Performances of two calcium/sodium–tailing systems The calcium/sodium–tailing composite was prepared by simply mixing tailing with solid calcium oxide and sodium hydroxide via mass ratio 1:4 and 1:2 respectively. The investigation procedures were the same as described in the section ‘Performances of two sole tailing systems’. Performance of two tailing–calcium/sodium–ferric systems The Cu/Fe tailing was put into a 250 mL Erlenmeyer in quantities of 1,000 mg/L and shaken at speed 150 rpm for 30 min. The pH after reaction was then adjusted to 10.0 by calcium/sodium alkali, right before adding 1 mg Fe3þ (ferric trichloride and ferric sulfate) into the suspensions and stirred at 300 rpm for 1 min, 60 rpm for 5 min, 45 rpm

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Application of three tailing-based composites in treating comprehensive electroplating wastewater

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for 5 min and 25 rpm for 5 min in steps. The supernatant for analysis was prepared as in the section ‘Performances of two sole tailing systems’. Effect of reaction time on performances The Cu/Fe tailing was put into a 250 mL Erlenmeyer in quantities of 1,000 mg/L and shaken at speed 150 rpm for 30, 60, 90, 120 and 240 min respectively to investigate effect of reaction time on performances. The pH was then adjusted to 10.0 by calcium oxide. The supernatant for analysis was prepared as in the section ‘Performances of two sole tailing systems’. Effect of pH on performances The Cu/Fe tailing was put into a 250 mL Erlenmeyer in quantities of 1,000 mg/L and shaken at speed 150 rpm for 30 min. The pH was then adjusted to 1.2, 4.3, 6.9, 8.7 and 10.0 respectively with calcium oxide. The supernatant for analysis was prepared as in the section ‘Performances of two sole tailing systems’. Performance stability and economic evaluation Three parallel tests were carried out to determine the stability and to analyze economic aspects of the calcium/sodium–Cu/ Fe tailing system. Sludge production of each test was determined by measuring suspended solids after reaction. The other investigation procedures were the same as described in the section ‘Performances of two sole tailing systems’.

Figure 1

T-Cu and T-Ni removal under different dosage and kinds of tailings.

Performance of two calcium-tailing systems Figure 2(a) shows the removal rates for T-Cu, T-Ni and COD in two calcium–tailing systems. The removal rates for T-Cu, T-Ni and COD were greater than 80% after 6 hours’ reaction; while the Cu/Fe tailing performed best in removing T-Cu, T-Ni and COD. Final pH values after reaction were also detected (Figure 2(b)), which were lower than the initial ones due to sedimentation of copper and nickel. Tailings combined with the hydrolysis of calcium/sodium were beneﬁcial to transport copper and nickel from aqueous solution to complex precipitation. COD removal might be attributed to oxidation by oxide contents in the tailings, which would perform better in acid conditions.

RESULTS AND DISCUSSION

Performance of two tailing–calcium–ferric systems

Performances of two sole tailing systems

Calcium, ferric trichloride and ferric sulfate were dosed to determine effects of coagulants on heavy metal removal. The average results are illustrated in Table 2. The calcium system could provide enough OH (hydroxyl radical) to remove T-Cu, T-Ni and COD from aqueous solution by sedimentation. The particulate matter in solution lost stability due to double layer compression during calcium hydrolysis and ferric ionization. It explained the fact that the calcium system performed similarly to the ferric trichloride and ferric sulfate systems under basic conditions.

The experiments were performed at room temperature in order to reduce differences caused by thermodynamic conditions. Tailing dosage was in quantities of 20, 100 and 1,000 mg/L respectively. Results are demonstrated in Figure 1. No signiﬁcant pollutant removal was observed for both of the sole tailing systems. Figure 1 indicates that both of the two sole tailing systems barely have any capacity in terms of T-Ni and T-Cu removal from the acid–nickel–copper electroplating wastewater, where the Cu/Fe tailing system performed slightly better than the Mn system when the dosage was low (20 mg/L). The remaining COD was 593.4 and 478.6 mg/L after reaction with 1,000 mg/L Mn tailing and Cu/Fe tailing respectively.

Effect of reaction time Dosage amount of the Cu/Fe tailing was 1.0 g/L. Time proﬁles for efﬂuent T-Cu and T-Ni indicated that removal of

H. Liu et al.

Figure 2

Table 2

Application of three tailing-based composites in treating comprehensive electroplating wastewater

Average efﬂuent T-Cu and T-Ni concentrations of different tailing systems

Efﬂuent T-Cu (mg/L)

Efﬂuent T-Ni (mg/L)

9.9

0.93

0.27

Tailing–calcium–ferric (Cl)

10.2

0.88

0.31

Tailing–calcium–ferric (sulfate)

10.2

0.90

0.35

Tailing–calcium

Initial pH

metals reached saturation within 30 min: the adsorption quantity did not increase, or even slightly decreased when the contact time increased from 30 to 120 min (Figure 3(a)). Time profile for effluent COD indicated that most of the COD was removed within the first 90 min after reaction. The COD removal rate increased very slowly with increasing contact time after 90 min: the effluent COD was only

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(a) Removal rates for T-Cu, T-Ni, and COD, (b) changes of pH after reaction.

Tailing Systems

Figure 3

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29 mg/L smaller when the contact time increased from 90 min to 4 hours (Figure 3(b)). Effect of pH Experiments were carried out with different initial pH values ranging from 1.0 to 10.0 using the calcium/sodium– Cu/Fe tailing system. The reaction time was set to be 30 min, and the Cu/Fe tailing dosage amount was 1.0 g/L. The removal rates for heavy metals were higher when the initial pH increased (Figure 4(a)). Increase of pH would release more hydroxyl ion in the system, which helped transform more heavy metals into chemical complexes such as copper hydroxide and nickel hydroxide (Ye et al. ). Figure 4(b) demonstrates that the effect of pH on COD removal was similar to that on heavy metals, which might be trigged by oxide contents in the tailings that performed alkaline oxidation.

(a) Effect of reaction time on efﬂuent T-Cu and T-Ni, (b) effect of reaction time on efﬂuent COD.

Figure 4

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Application of three tailing-based composites in treating comprehensive electroplating wastewater

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(a) Effect of pH on efﬂuent T-Cu and T-Ni, (b) effect of pH on efﬂuent COD.

Effect of initial concentrations

Adsorption isotherms

Effect of initial concentration on T-Cu and T-Ni removal was investigated using the calcium/sodium–Cu/Fe tailing system with optimized mass ratio of 4:2:1; the pH was set as 10.0–11.0. Eight parallel tests with initial concentrations of 1, 10, 20, and 50 mg/L for Cu(II) and 5, 20, 50, 100 mg/L for Ni(II) were carried out in eight 200 mL Erlenmeyers and shaken at the speed of 150 rpm for 300 min. T-Cu and T-Ni contents after reaction under different initial concentrations are demonstrated in Figures 5(a) and 5(b). Removal rates of more than 90% were achieved under all the testing initial concentrations. Higher initial concentrations were favorable to T-Cu and T-Ni removal but with insigniﬁcant effects. For example, by increasing initial Cu(II) concentration from 1 to 50 mg/L, the T-Cu removal rates increased from 90.91 to 99.08%; while the removal rates for T-Ni increased from 93.70 to 99.79% when the initial concentration increased from 5 to 100 mg/L.

The isotherms of the novel composite material for T-Cu and T-Ni adsorption were obtained using the calcium/sodium– Cu/Fe tailing system with the mass ratio of 4:2:1. Adsorption tests with different initial contents of heavy metals, i.e. 1, 2, 5, 10 mg/L for Cu (II) and 5, 15, 30 mg/L for Ni(II), were carried out in 200 mL Erlenmeyers and shaken at the speed of 150 rpm. Sampling intervals for all the adsorption tests were 30 min until the saturation was reached (metal contents in the supernatant came to a stable level). The pH was set as 10.0–11.0. Adsorption proﬁles indicated that the isotherms for Cu(II) and Ni(II) adsorption ﬁtted the Langmuir and Freundlich models respectively with R 2 values as high as 0.92 (Figures 6(a) and 6(b)). The results indicated that the tailing composite removed Cu(II) and Ni(II) with different mechanisms. The removal of Cu(II) was mainly caused by monolayer adsorption, where the activity points distributed

Figure 5

(a) Cu removal rates under different initial concentrations, (b) Ni removal rates under different initial concentrations.

Figure 6

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Application of three tailing-based composites in treating comprehensive electroplating wastewater

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(a) Langmuir ﬁtted isotherm for Cu(II) adsorption by the composite, (b) Freundlich ﬁtted isotherm for Ni(II) adsorption by the composite.

evenly on the surface of the adsorbent, while the removal of Ni(II) was mainly caused by multilayer adsorption, where the activity points distributed on sites of varied afﬁnities. It also implied that the adsorption capacity of Cu(II) and Ni(II) on the tailing composite could be diverse mainly due to different hydrated ionic radius of the two heavy metals since their electric charges were identical. Other parameters and constants for the ﬁtted isotherms are listed in Table 3. The Langmuir plots (Figure 6(a)) were based on the isotherm equation

(mg/g); Kf and n were Freundlich constants (Ajmal et al. ). The value of n was 0.71, not falling in the range of 1–10, indicating also poor physical adsorption between Ni(II) and composites (Oubagaranadin & Murthy ). Values for qm and Kf were on the higher sides although b and n values predicted otherwise, which implied signiﬁcant involvement of chemical adsorption and/or chemical reaction in the process.

Ce =qe ¼ 1=ðbqm Þ þ ð1=qm ÞCe

Three parallel tests were carried out to investigate performance stability of the Cu/Fe tailing–calcium–sodium composite with mass ratio 4:2:1. Dosage of the composite was 7 g/L and the shaking time was 30 min. Results are illustrated in Table 4. It showed that performance of the composite was fairly stable with maximum RSD (relative standard deviation) less than 12% within the three tests. The removal efﬁciency of COD is 71.8%. Cost for sludge disposal was a signiﬁcant part of total operation cost in electroplating wastewater treatment (Sousa et al. ). Figures 7(a) and 7(b) showed that the calcium/sodium system with mass ratio 2:1 produced 10% less sludge than that of the sole calcium system.

(1)

where Ce/qe was metal uptake per unit weight of adsorbent (mg/g); Ce was the equilibrium concentration of metal (mg/L); qm and b were Langmuir constants relating to adsorption capacity and adsorption energy respectively. The value of b was 35.3, indicating a weak physical afﬁnity between the Cu(II) and composites. The Freundlich plots (Figure 6(b)) were obtained from the isotherm equation LogðCe =qe Þ ¼ log Kf þ 1=ðn log Ce Þ

(2)

where Ce was the equilibrium concentration (mg/L); Ce/qe was the amount adsorbed per unit weight of adsorbent Table 3

Performance stability and economy evaluation

Table 4

Results of parallel experiments

Efﬂuent

Isotherm models, constants and statistical parameters for Cu(II) and Ni(II) adsorption

Heavy metals

Isotherm models

Items

Constants

Values

Cu(II)

Langmuir

0.92

qm b

154.56 35.25117

Ni(II)

Freundlich

0.92

Kf n

24,620.1 0.72

Initial

Reaction

T-Cu

T-Ni

COD

(mg/L)

Parallel test one

1.1

11.8

0.35

0.40

200.9

Parallel test two

1.1

11.6

0.28

0.40

172.4

0.34

212.9

Parallel test three

1.2

11.6

RSD (%)

0.51 12

10.6

H. Liu et al.

Figure 7

Application of three tailing-based composites in treating comprehensive electroplating wastewater

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(a) Weight of sludge produced by the calcium system and the calcium/sodium system; None: control (suspended solids in wastewater), (b) weight percentage of sludge produced by each system to the total sludge produced by three systems; None: control (suspended solids in wastewater).

Table 5

Brief economic comparison between the BAU process and the TCP process

Cost for oxidant

Cost for alkali

Cost for sludge

(Yuan/m3-treated

treatment (Yuan/m3-

Cost for tailing (Yuan/m3-treated

chemical (Yuan/m3-

Efﬂuent

Processes

water)

treated water)

water)

treated water)

(mg/L)

BAU

0.63

13.24

6.37

20.24

T-Cu: 0.57 T-Ni: 1.54

TCP

4.00

5.60

0.06 (for transportation only)

9.66

T-Cu: 0.88 T-Ni: 0.37

Oxidant þ alkali þ sedimentation is the business-asusual process (BAU) to remove heavy metals from electroplating wastewater, where hydrogen peroxide is normally used as the oxidant to break complexing between metals and some organic groups (Gupta ). Researchers also proposed activated carbon as an alternative process, but it is not acceptable at the industry scale due to economic reasons (Akhtar et al. ). A brief economic comparison between the BAU process and the optimized Cu/Fe tailing–calcium–sodium composite process (TCP) is shown in Table 5, where only cost for chemicals is included. Other costs such as electricity and labour costs are at the same level for both processes. The dosage amounts were obtained by parallel tests, while prices for chemicals were provided by the studied electroplating wastewater plant. Table 5 indicates that the TCP process was more cost-attractive than the BAU process due to three reasons: (1) the tailing contains manganese dioxide, which served as alternative oxidant in the BAU process; (2) the tailing contains alkali substances such as magnesium oxide, which replenished part of the alkali pool needed in the BAU process; (3) the

Total cost for dosing

tailing is a mineral waste, and only cost for transportation is needed.

CONCLUSIONS Heavy metals and COD are major challenging pollutants for most electroplating wastewater treatment plants. The composite material proposed in the work is simple to prepare by mixing calcium/sodium compounds and tailings, and proved to be both economic and efficient in combating comprehensive electroplating pollution. The calcium/ sodium–Cu/Fe tailing system was relatively stable with less sludge production, which increased the potential of applying the composite in industrial-scale electroplating wastewater plants. The composite showed poor physical adsorption but strong chemical adsorption and/or reaction; thus profiling of the tailing such as morphological studies would help improve treatment efficiency of the wastewater; while stoichiometric characteristics and chemical coordination

H. Liu et al.

Application of three tailing-based composites in treating comprehensive electroplating wastewater

between tailing and alkali would be the focus for further research.

ACKNOWLEDGEMENT This study was ﬁnancially supported by the Natural Science Foundation of China (grant no. 21206092).

REFERENCES Ajmal, M., Rao, R. A. K., Ahmad, R. & Ahmad, J.  Adsorption studies on Citrus reticulata (fruit peel of orange): removal and recovery of Ni (II) from electroplating wastewater. Journal of Hazardous Materials 79 (1–2), 117–131. Ajmal, M., Rao, R. A. K., Ahmad, R., Ahmad, J. & Rao, L. A. K.  Removal and recovery of heavy metals from electroplating wastewater by using Kyanite as an adsorbent. Journal of Hazardous Materials 87 (1–3), 127–137. Akhtar, N., Iqbal, J. & Iqbal, M.  Removal and recovery of nickel (II) from aqueous solution by loofa spongeimmobilized biomass of Chlorella sorokiniana: characterization studies. Journal of Hazardous Materials 108 (1–2), 85–94. Anirudhan, T. S. & Suchithra, P. S.  Humic acid-immobilized polymer/bentonite composite as an adsorbent for the removal of copper(II) ions from aqueous solutions and electroplating industry wastewater. Journal of Industrial and Engineering Chemistry 16 (1), 130–139. APHA  Standard Methods for the Examination of Water and Wastewater, 20th edn. American Public Health Association/ American Water Works Association/Water Environment Federation, Washington, DC, USA. Congeevaram, S., Dhanarani, S., Park, J., Dexilin, M. & Thamaraiselvi, K.  Biosorption of chromium and nickel by heavy metal resistant fungal and bacterial isolates. Journal of Hazardous Materials 146 (1–2), 270–277. Gupta, V. K.  Equilibrium uptake, sorption dynamics, process development, and column operations for the removal of copper and nickel from aqueous solution and wastewater using activated slag, a low-cost adsorbent. Industrial & Engineering Chemistry Research 37 (1), 192–202.

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Hawari, A. H. & Mulligan, C. N.  Biosorption of lead(II), cadmium(II), copper(II) and nickel(II) by anaerobic granular biomass. Bioresource Technology 97 (4), 692–700. Kadirvelu, K., Thamaraiselvi, K. & Namasivayam, C.  Adsorption of nickel (II) from aqueous solution onto activated carbon prepared from coirpith. Separation and Puriﬁcation Technology 24 (3), 497–505. Kurniawan, T. A., Chan, G. Y. S., Lo, W. H. & Babel, S.  Comparisons of low-cost adsorbents for treating wastewaters laden with heavy metals. Science of the Total Environment 366 (2–3), 409–426. Oubagaranadin, J. U. K. & Murthy, Z. V. P.  Isotherm modeling and batch adsorber design for the adsorption of Cu (II) on a clay containing montmorillonite. Applied Clay Science 50 (3), 409–413. Padmavathy, V., Vasudevan, P. & Dhingra, S. C.  Biosorption of nickel (II) ions on baker’s yeast. Process Biochemistry 38 (10), 1389–1395. Sciban, M., Radetic, B., Kevresan, D. & Klasnja, M.  Adsorption of heavy metals from electroplating wastewater by wood sawdust. Bioresource Technology 98 (2), 402–409. Sheng, P. X., Ting, Y. P., Chen, J. P. & Hong, L.  Sorption of lead, copper, cadmium, zinc, and nickel by marine algal biomass: characterization of biosorptive capacity and investigation of mechanisms. Journal of Colloid and Interface Science 275 (1), 131–141. Sousa, F. W., Sousa, M. J., Oliveira, I. R. N., Oliveira, A. G., Cavalcante, R. M., Fechine, P. B. A., Neto, V. O. S., de Keukeleire, D. & Nascimento, R. F.  Evaluation of a lowcost adsorbent for removal of toxic metal ions from wastewater of an electroplating factory. Journal of Environmental Management 90 (11), 3340–3344. Vengris, T., Binkiene, R. & Sveikauskaite, A.  Nickel, copper and zinc removal from waste water by a modiﬁed clay sorbent. Applied Clay Science 18 (3–4), 183–190. Villaescusa, I., Fiol, N., Martinez, M., Miralles, N., Poch, J. & Serarols, J.  Removal of copper and nickel ions from aqueous solutions by grape stalks wastes. Water Research 38 (4), 992–1002. Yavuz, O., Altunkaynak, Y. & Guzel, F.  Removal of copper, nickel, cobalt and manganese from aqueous solution by kaolinite. Water Research 37 (4), 948–952. Ye, J. S., Yin, H., Mai, B. X., Peng, H., Qin, H. M., He, B. Y. & Zhang, N.  Biosorption of chromium from aqueous solution and electroplating wastewater using mixture of Candida lipolytica and dewatered sewage sludge. Bioresource Technology 101 (11), 3893–3902.

First received 27 November 2013; accepted in revised form 7 April 2014. Available online 21 April 2014

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Reduction of hexavalent chromium: photocatalysis and photochemistry and their application in wastewater remediation Tiele Caprioli Machado, Marla Azário Lansarin and Natália Matte

ABSTRACT Hexavalent chromium present in wastewater discharge of galvanic industries is toxic to most microorganisms and potentially harmful to human health. This work examines the photochemical reduction of Cr(VI) with ethanol under ultraviolet (UV) and visible radiation, and photocatalytic reduction of Cr(VI) with TiO2 in the presence of ethanol under UV radiation. By means of different experimental designs, this study investigates the inﬂuence of the initial pH, ethanol amount, catalyst

Tiele Caprioli Machado (corresponding author) Marla Azário Lansarin Natália Matte Chemical Engineering Department, Federal University of Rio Grande do Sul Rua Engenheiro Luiz Englert s/n – CEP: 90040-040 – Porto Alegre, RS – Brazil E-mail: tiele@enq.ufrgs.br

concentration and initial Cr(VI) concentration on total Cr(VI) reduction. The results obtained showed that photochemistry with ethanol under UV radiation (96.10%) was more efﬁcient than photochemistry with ethanol under visible light (48.07%). Furthermore, photocatalysis with TiO2 in the presence of ethanol under UV radiation showed high values of total Cr(VI) reduction: 94.15%, under the optimal conditions established by the experimental design. Finally, experiments were carried out with wastewater discharge from an electroplating plant in its original concentration, and higher values of total Cr(VI) reduction were observed. Key words

| hexavalent chromium, photocatalysis, photochemistry, wastewater

INTRODUCTION Hexavalent chromium – Cr(VI) – is toxic to most microorganisms and potentially harmful to human health. In the environment, Cr(VI), which is more stable than trivalent chromium, is highly soluble, making the contamination of groundwater and drinking water sources a possibility. The hexavalent form of chromium is 100 times more toxic than the trivalent (Mohapatra et al. ; Das et al. ; Xu et al. ; Yang & Chan ). A possible source of chromium is the improper discharge of wastewaters from electroplating plants. The conventional treatment for these wastewaters is reduction with sodium sulﬁte or metabisulﬁte. This method, however, requires an excessive amount of chemicals to ensure that the complete conversion of Cr(VI) and the formation of sludge or the release of sulfur dioxide occurs. Studies have been directed to developing cleaner technologies capable of dealing with wastewater containing chromium. Among these are heterogeneous photocatalysis and photochemistry with alcohols (Chakrabarti et al. ). Literature shows few studies on the photochemical reduction of Cr(VI) by alcohols (Mytych et al. ), doi: 10.2166/wst.2014.193

phenol and its halogen derivatives (Mytych & Stasicka ), and glycerol (Yurkow et al. ). In these studies, only Mytych et al. () used ethanol in the photochemical reduction reaction of Cr(VI), and the work is limited to the effects of the alcohol concentration. Also, a comparison between photochemical reduction reactions of Cr(VI) with ethanol under ultraviolet (UV) radiation and visible light was not found in literature. Furthermore, there are several studies on the photocatalytic reduction of Cr(VI) with TiO2 (Ku & Jung ; Schrank et al. ; Cho et al. ; Tuprakay & Liengcharernsit ; Xu et al. ; Jiang et al. ). However, the ethanol effect on the photocatalytic reduction of Cr(VI) with TiO2 under UV radiation was studied only by Ku & Jung () and in this study the authors analyzed the ethanol effect, leaving out other variables that affect the reaction rate. There is no systematic and comparative study that analyzes the effect of the solution’s pH, catalyst concentration, initial Cr(VI) concentration and ethanol amount on the photocatalytic reduction of Cr(VI) with TiO2 in the presence of ethanol under UV radiation. The same happens with photochemical

T. C. Machado et al.

Photocatalysis and photochemistry and their application in wastewater remediation

reduction of Cr(VI) with ethanol under UV and visible radiation. Furthermore, the application of sustainable treatments such as photochemistry and heterogeneous photocatalysis to remediate Cr(VI) present in wastewater discharges of electroplating industries has not been found in the literature. This work aims to ﬁll this information gap and, thus, the reduction of Cr(VI) by photochemistry with alcohols under UV and visible radiation, and photocatalysis with TiO2 in the presence of ethanol under UV radiation were investigated. To evaluate Cr(VI) reduction by photochemistry with ethanol, two experimental designs (UV and visible) were used in which the inﬂuence of initial pH, ethanol amount and initial Cr(VI) concentration were investigated. Subsequently, tests of photocatalytic reduction of Cr(VI) with TiO2 in the presence of ethanol under UV radiation were carried out and the effects of the initial pH, ethanol amount, catalyst concentration and initial Cr(VI) concentration were evaluated through an experimental design. Finally, experiments were conducted using raw wastewater from an electroplating plant.

MATERIALS AND METHODS Materials Deionized and distilled water and potassium dichromate (Fmaia), as it was received, were used to prepare the solutions. Absolute ethyl alcohol (Nuclear) was used in the reactions of photochemical reduction of Cr(VI), and a titanium dioxide (TiO2) catalyst from Degussa P-25, 80% anatase and 20% rutile with a specific area of 50 m2 g 1 and an average particle size of 20 nm was used in the reactions of photocatalytic reduction of Cr(VI). Experiments of Cr(VI) reduction The experiments were carried out in a batch reactor irradiated with a mercury vapor lamp (Philips HPL-N 125 Watt) in its original shape for visible light, or modified to emit in the UV region by removing the inner covering of the bulb. Agitation of the reaction medium was conducted by a magnetic stirrer, and the temperature was monitored by using a K-type thermocouple connected to a digital display. The irradiation chamber consisted of a wooden box, internally lined with aluminum foil, and a syringe-catheter system was used to collect liquid samples at specific intervals. Lamp radiation was adjusted by changing the distance between the lamp and the solution, and was measured at

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the beginning of each test using a radiometer (Cole-Parmer Instrument, Series 9811 Radiometer) for UV radiation and a solarimeter (Kimo® Instruments) for visible radiation. The synthetic Cr(VI) solution was prepared by using potassium dichromate (K2Cr2O7), at concentrations ranging from 10 to 50 mg L 1, and kept under vigorous magnetic stirring for approximately 1 h. For the photochemical reduction reactions with ethanol, a known volume of alcohol was added to 150 mL of the aqueous Cr(VI) solution and the solution’s pH was adjusted. These reactions, under UV and visible radiation, occurred in 1 h, and the progress of the photochemical reduction reaction of Cr(VI) was monitored by collecting samples at fixed intervals (0, 15, 30, 45 and 60 min). Subsequently, the Cr(VI) content was analyzed quantitatively by measuring the absorbance at 348 nm using a Cary 100 (Varian) Spectrophotometer (Mohapatra et al. ; Das et al. ; Nasrallah et al. ). A calibration curve was used to relate the Cr(VI) concentration in the samples collected during the tests to its absorbance at a wavelength of maximum absorption (348 nm), according to the Lambert–Beer Law. For the photocatalytic reduction reactions of Cr(VI) with TiO2 in the presence of ethanol under UV radiation, catalyst powder was added to 150 mL of aqueous Cr(VI) solution to produce a suspension of known concentration, and the solution’s pH was adjusted. The reduction tests were divided into two steps: one in the dark (1 h), in which adsorption– desorption equilibrium of the metal on the catalyst surface occurs without radiation, and the other (1 h) with UV radiation. The progress of the photocatalytic reduction reaction of Cr(VI) was monitored by collecting samples at fixed intervals (0, 15, 30, 45 and 60 min). Subsequently, the samples were centrifuged for 15 min, filtered with an ester mixture membrane (0.2 μm) and stored in amber glass bottles, and then analyzed in a spectrophotometer. For the tests carried out with raw wastewater, samples were diluted with distilled and deionized water to obtain the same chromium concentration as under the best conditions of the experimental design (20 mg L 1). Experiments were also carried out with raw wastewater in its original Cr(VI) concentration, which was 1,100 mg L 1. These tests followed the same methodology as used for the synthetic Cr(VI) solution. Methodology of response surface The statistical model of a circumscribed central composite design was used in order to optimize ethanol amount, initial pH of the solution and initial Cr(VI) concentration for

T. C. Machado et al.

Photocatalysis and photochemistry and their application in wastewater remediation

photochemical reduction reactions of Cr(VI) with ethanol under UV and visible radiation. This model consists of a factorial design with 2n (n ¼ 3) experiments, six star points and three central points, resulting in a total of 17 experiments for each radiation. For the photocatalytic reduction reactions of Cr(VI) with TiO2 in the presence of ethanol under UV radiation, the independent variables were ethanol amount, catalyst concentration, initial pH of the solution, and initial Cr(VI) concentration. This model of a factorial design with 2n (n ¼ 4) experiments, eight star points and three central points, resulted in a total of 27 experiments. The response factor, for all of the experimental designs, was deﬁned as the total Cr(VI) reduction after 1 h of reaction. Data were analyzed using the Statistica 10 software and the model was statistically validated using analysis of variance (conﬁdence interval 95%).

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radiation. The numbers in parentheses represent the coded variables, and the response factor (Y ) was defined as the percentage of total Cr(VI) reduction after 1 h of reaction. The polynomial (Equations (1) and (2)) obtained from the multivariable analysis showed a correlation coefficient (R 2) of 0.9431 for reactions using UV radiation, and 0.9527 for reactions using visible radiation, indicating a good fit between the experimental data and the statistical model. In the polynomial equations, the values in parentheses represent standard deviation for each coefficient and x, y and z represent the coded values of the initial pH, initial Cr(VI) concentration and ethanol amount, respectively. YUV (%) ¼ 72:3ð±3:8Þ 25:3xð±2:6Þ 11:5x2 ð±2:6Þ 5:8yð±2:6Þ þ 17:1zð±2:6Þ 9:0z2 ð±2:6Þ 11:5x:zð±3:4Þ

(1)

YVisible (%) ¼ 24:0ð±1:3Þ 10:4xð±1:1Þ 3:4x2 ð±1:1Þ

RESULTS AND DISCUSSION

2:5yð±1:1Þ þ 10:4zð±1:1Þ 6:1x:zð±1:4Þ

Optimization of the photochemical reduction reaction of Cr(VI) with ethanol Table 1 shows the experiments of the photochemical reduction of Cr(VI) with ethyl alcohol under UV and visible

(2)

The initial pH of the solution and the ethanol amount are the main variables affecting the reaction, with linear and quadratic interactions. In addition, for both

Table 1

Tests

Initial pH of the solution

Initial Cr(VI) concentration (mg L 1)

Ethanol amount (%) (vethanol/vtotal )

Total reduction (%) (Yexp.) UV radiation

Total reduction (%) (Yexp.) visible radiation

2 ( 1)

20 ( 1)

1.5 ( 1)

55.51

15.65

3.5 (1)

20 ( 1)

1.5 ( 1)

28.89

4.74

2 ( 1)

40 (1)

1.5 ( 1)

44.24

13.09

3.5 (1)

40 (1)

1.5 ( 1)

19.35

5.61

2 ( 1)

20 ( 1)

5 (1)

96.10

48.07

3.5 (1)

20 ( 1)

5 (1)

34.79

13.68

2 ( 1)

40 (1)

5 (1)

94.88

39.91

3.5(1)

40 (1)

5 (1)

12.72

7.42

1.5 ( 1.68)

30 (0)

3.25 (0)

89.55

34.02

4 (1.68)

30 (0)

3.25 (0)

0.00

0.45

2.75 (0)

13 ( 1.68)

3.25 (0)

86.03

29.78

2.75 (0)

47 (1.68)

3.25 (0)

64.97

18.84

2.75 (0)

30 (0)

0.31 ( 1.68)

9.29

2.49

2.75 (0)

30 (0)

6.2 (1.68)

94.48

45.22

2.75 (0)

30 (0)

3.25 (0)

71.36

28.26

2.75 (0)

30 (0)

3.25 (0)

73.98

27.44

2.75 (0)

30 (0)

3.25 (0)

71.65

27.06

Factorial experimental design for photochemical reduction reaction of Cr(VI) with ethanol under UV and visible radiation

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Photocatalysis and photochemistry and their application in wastewater remediation

photochemical reactions (UV and visible), the coefficients of the linear effects of pH and Cr(VI) concentration were negative, indicating that the lower the initial pH of the solution and the initial Cr(VI) concentration, the higher the total Cr(VI) reduction. The coefficient for the linear effect of ethanol amount was positive, so the greater the ethanol amount, the higher the total Cr(VI) reduction. Moreover, the intensity of the coefficients of the polynomial was greater, in all cases, under UV radiation. This behavior can be seen in Table 1, in which test 5, for reactions under UV and visible radiation, with lower initial pH values and Cr(VI) concentration and high ethanol amount, presented the highest total Cr(VI) reduction of 96.10 and 48.07%, respectively. Moreover, it was observed that the photochemical reactions with ethanol under UV radiation obtained higher values of Cr(VI) reduction than the photochemical reactions with ethanol under visible light. The effect of initial pH has been associated with higher susceptibility of HCrO 4 to reduction and chromate(VI) ester formation, a necessary condition for redox or photoredox reactions. Equations (3)–(5) show the Cr(VI) species distribution in aqueous solution as a function of the solution’s pH. The neutral chromic acid molecule, H2CrO4, is the predominant species for pH levels below 2.0, while 2 concentrations of negatively charged HCrO and 4 , CrO4 2 Cr2O7 species are predominant for pH levels above 2.0 (Ku & Jung ); and Equations (6)–(10) show the mechanism of the Cr(VI) reduction reaction by alcohols (Mytych et al. ).

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The total Cr(VI) reduction is proportional to the amount of electrons supplied by the alcohol (Equation (6)). Therefore, the greater the ethanol amount, the more electrons are provided to reduce the initial Cr(VI) concentration, increasing total Cr(VI) reduction. Conversely, maintaining the ethanol amount at a ﬁxed level and increasing the initial Cr(VI) concentration will decrease total Cr(VI) reduction. In addition, other tests were conducted in the absence of UV or visible light to verify Cr(VI) reduction in the presence of alcohol. No Cr(VI) reduction was observed in 2 h, indicating that Cr(VI) reduction occurs only in the presence of light; thus it is considered a photochemical reaction. Similar behavior was observed by Mytych et al. (). To quantify the contribution of photolysis, experiments in which only the Cr(VI) solution was subjected to UV radiation/visible radiation were also carried out. The reaction under UV showed no Cr(VI) reduction, while the reaction under visible light showed a total Cr(VI) reduction of 0.4%. Thus, the contribution of photolysis can be negligible (Mohapatra et al. ; Chakrabarti et al. ).

Optimization of the photocatalytic reduction reaction of Cr(VI) with TiO2 in the presence of ethanol under UV radiation

H2 CrO4 ↔ Hþ þ HCrO 4

(3)

2 þ HCrO 4 ↔ H þ CrO4

(4)

2 2HCrO 4 ↔ Cr2 O7 þ H2 O

(5)

HCrO 4 þ ROH ↔ ROCrO3 þ H2 O

(6)

ROCrO 3 þ hv ! Cr(IV) þ aldehyde=ketone

(7)

Cr(IV) þ Cr(VI) ! 2Cr(V)

(8)

For the photocatalytic reduction reactions of Cr(VI) with TiO2 in the presence of ethanol under UV radiation, the variables studied were ethanol amount, catalyst concentration, initial pH of the solution, and initial Cr(VI) concentration. The experimental design and results are shown in Table 2. The numbers in parentheses represent the coded variables, and the response factor (Y) was deﬁned as the percentage of total Cr(VI) reduction after 1 h of reaction. Equation (11) shows the polynomial obtained from the multivariable analysis. The values in parentheses represent standard deviation for each coefﬁcient and x, y, z and w represent the coded values of the initial pH, initial Cr(VI) concentration, ethanol amount and TiO2 concentration respectively.

3Cr(V) ! 2Cr(VI) þ Cr(III)

(9)

Y(%) ¼ 30:7ð±1:4Þ 12:5xð±0:8Þ þ 6:7x2 ð±0:8Þ þ 12:3wð±0:8Þ 17:3yð±0:8Þ þ 5:9y2 ð±0:8Þ

Cr(V) þ ROH ! Cr(III) þ aldehyde=ketone

(10)

þ 1:8zð±0:8Þ 3:3x:wð±1:0Þ 3:7w:yð±1:0Þ

(11)

Table 2

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Photocatalysis and photochemistry and their application in wastewater remediation

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Factorial experimental design for photocatalytic reduction reaction of Cr(VI) with TiO2 in the presence of ethanol under UV radiation

Tests

Initial pH of the solution

Catalyst concentration (g L 1)

Initial Cr(VI) concentration (mg L 1)

Ethanol amount (%) (vethanol/vtotal )

Total reduction (%) (Yexp.)

3 ( 1)

0.8 ( 1)

20 ( 1)

0.15 ( 1)

46.92

6 (1)

0.8 ( 1)

20 ( 1)

0.15 ( 1)

32.81

3 ( 1)

2.0 (1)

20 ( 1)

0.15 ( 1)

84.85

6 (1)

2.0 (1)

20 ( 1)

0.15 ( 1)

62.18

3 ( 1)

0.8 ( 1)

40 (1)

0.15 ( 1)

28.57

6 (1)

0.8 ( 1)

40 (1)

0.15 ( 1)

8.91

3 ( 1)

2.0 (1)

40 (1)

0.15 ( 1)

49.69

6 (1)

2.0 (1)

40 (1)

0.15 ( 1)

21.73

3 ( 1)

0.8 ( 1)

20 ( 1)

0.35 (1)

56.16

6 (1)

0.8 ( 1)

20 ( 1)

0.35 (1)

39.66

3 ( 1)

2.0 (1)

20 ( 1)

0.35 (1)

92.97

6 (1)

2.0 (1)

20 ( 1)

0.35 (1)

60.35

3 ( 1)

0.8 ( 1)

40 (1)

0.35 (1)

26.24

6 (1)

0.8 ( 1)

40 (1)

0.35 (1)

14.31

3 ( 1)

2.0 (1)

40 (1)

0.35 (1)

52.44

6 (1)

2.0 (1)

40 (1)

0.35 (1)

20.05

1.5 ( 2)

1.4 (0)

30 (0)

0.25 (0)

87.37

7.5 (2)

1.4 (0)

30 (0)

0.25 (0)

26.51

4.5 (0)

0.2 ( 2)

30 (0)

0.25 (0)

6.38

4.5 (0)

2.6 (2)

30 (0)

0.25 (0)

59.08

4.5 (0)

1.4 (0)

10 ( 2)

0.25 (0)

94.15

4.5 (0)

1.4 (0)

50 (2)

0.25 (0)

13.50

4.5 (0)

1.4 (0)

30 (0)

0.05 ( 2)

24.62

4.5 (0)

1.4 (0)

30 (0)

0.45 (2)

33.45

4.5 (0)

1.4 (0)

30 (0)

0.25 (0)

29.00

4.5 (0)

1.4 (0)

30 (0)

0.25 (0)

30.54

4.5 (0)

1.4 (0)

30 (0)

0.25 (0)

29.51

The polynomial presents a correlation coefficient (R²) of 0.9823 and indicates that the initial pH of the solution and the initial Cr(VI) concentration are the main variables that affect the reaction, with linear and quadratic interactions. The linear coefficients of the effects of the initial pH and the Cr(VI) concentration were negative, indicating that the lower the initial pH of the solution and the initial Cr(VI) concentration, the higher the total Cr(VI) reduction. The coefficients for the linear effects of the TiO2 concentration and the ethanol amount are positive, indicating that the higher the TiO2 concentration and the ethanol amount, the higher the total Cr(VI) reduction. As can be seen in Table 2, test 21 showed the largest removal of Cr(VI) at 94.15%, with an initial

pH of 4.5, an initial Cr(VI) concentration of 10 mg L 1, a TiO2 concentration of 1.4 g L 1, and an ethanol amount of 0.25% (v/v). The initial pH of the solution is an important parameter, inﬂuencing reduction reactions of Cr(VI). The initial pH determines the charges to the catalyst surface as well as the ionization state of the substrate, which affects the amount adsorbed and hence the total reduction. The predominant species of Cr(VI) at pH 2 is HCrO 4 . Thus, an increase in the solution’s pH will shift the HCrO 4 concen2 tration to form CrO2 and Cr O species. As the point of 4 2 7 zero charge of TiO2 is pH 7.52 (Ku & Jung ), at pH values above the isoelectric point, the semiconductor surface becomes negative and repels dichromate ions (Ku &

T. C. Machado et al.

Photocatalysis and photochemistry and their application in wastewater remediation

Jung ; Mohapatra et al. ). Therefore, the higher the initial pH of the solution, the lower the Cr(VI) reduction. As observed, by decreasing the initial Cr(VI) concentration, an increase in the total reduction of this species takes place because at higher levels of the substrate’s initial concentration, absorbance of Cr(VI) solution is increased and a higher fraction of the UV radiation is intercepted before it reaches the catalyst surface, thereby diminishing the efﬁciency of the photocatalytic reaction (Ku & Jung ; Das et al. ; Chakrabarti et al. ). The catalyst, when illuminated with energy of wavelength equal to or greater than its band-gap energy, generates electron/hole pairs. These, in turn, reduce the Cr(VI) and oxidize ethanol, which acts as a sacriﬁcial electron donor providing electrons to the reaction medium and prevents the recombining of electron/hole pairs. Greater Cr(VI) reductions were observed with photocatalytic reactions of Cr(VI) with TiO2 in the presence of ethanol, demonstrating the synergistic effect between ethanol oxidation and Cr(VI) reduction, as can be seen in Figure 1. Experiments were conducted in the absence of UV or visible light to verify the adsorption of Cr(VI) on the catalyst surface. An adsorption of 14% was observed in 1 h, with an initial pH at 3.0, an initial Cr(VI) concentration at 20 mg L 1, catalyst concentrations of 2.0 g L 1, and ethanol amount at 0.35% (v/v). Tests with wastewater In order to verify the application of photochemistry with ethanol and heterogeneous photocatalysis in treating Cr(VI) in raw wastewater discharges from an electroplating

Water Science & Technology

Photocatalytic reduction of Cr(VI) with TiO2 under UV radiation in the presence and absence of ethanol (CTiO2 ¼ 2.0 g L 1, CCr(VI) ¼ 20 mg L 1, Vethanol ¼ 0.35% (v/v), pH ¼ 3.0, T ¼ 30 C, and 5.8 mW cm 2). W

2014

Tests of total Cr(VI) reduction with diluted wastewater

Source of Cr(VI)

70.1

plant, tests of total Cr(VI) reduction were initially carried out with diluted wastewater, according to the methodology mentioned above, using the conditions that yielded maximum reduction as indicated by the experimental design. The results obtained were similar to those obtained with the synthetic solution, and signiﬁcantly improved in the case of photocatalysis with TiO2, in the presence of ethanol under UV radiation, as observed in Table 3. Subsequently, tests of total Cr(VI) reduction were carried out with industrial wastewater in its original concentration (1,100 mg L 1), under other experimental conditions, in order to enable the application of these processes in electroplating industries. These tests were conducted to evaluate the ethanol amount and the reaction time required to reduce the Cr(VI) present in concentrated wastewater. The conditions tested for the synthetic solution in photochemistry with ethanol under UV radiation showing 96.10% removal were used to evaluate the ethanol amount and the reaction time. Results indicated the highest total Cr(VI) reduction to be 98.48%, considering an ethanol amount of 40% (v/v) and a reaction time of 6 h. Thus, these conditions were selected to run photochemical reduction reactions of Cr(VI) with ethanol under visible light and photocatalytic reduction reactions of Cr(VI) with TiO2, in the presence of ethanol, under UV radiation (Table 4). As can be observed in Table 4, all processes were successful in obtaining high removals of total Cr(VI), suggesting their potential to be used for industrial applications. Moreover, the most suitable process is photochemistry with ethanol under UV radiation. Although it does not present the largest total Cr(VI) reduction, it does

Table 3

Figure 1

Total Cr(VI)

solution

Reactions

reduction (%)

Synthetic (K2Cr2O7)

Photochemistry ethanol, UV

96.10

Diluted wastewater

Photochemistry ethanol, UV

95.10

Synthetic (K2Cr2O7)

Photochemistry ethanol, visible

48.07

Diluted wastewater

Photochemistry ethanol, visible

44.87

Synthetic (K2Cr2O7)

Photocatalysis TiO2, UV, with ethanol

92.97

Diluted wastewater

Photocatalysis TiO2, UV, with ethanol

99.62

Table 4

T. C. Machado et al.

Photocatalysis and photochemistry and their application in wastewater remediation

Tests of total Cr(VI) reduction with industrial wastewater in its original concentration

Total Cr(VI) Reactions

Experimental conditions

reduction (%)

Photochemistry ethanol, UV

CCr(VI) ¼ 1,100 mg L 1, pH ¼ 2.0, Vethanol ¼ 40%, 5.8 mW cm 2, 6 h of reaction

98.48

Photochemistry ethanol, visible

CCr(VI) ¼ 1,100 mg L 1, pH ¼ 2.0, Vethanol ¼ 40%, 250 W m 2, 6 h of reaction

89.74

Photocatalysis TiO2, UV, with ethanol

CCr(VI) ¼ 1,100 mg L 1, pH ¼ 2.0, CTiO2 ¼ 2.0 g L 1, Vethanol ¼ 40%, 5.8 mW cm 2, 6 h of reaction

99.21

not require catalyst separation processes and presents lower reagent costs, since ethanol is inexpensive, non-toxic and easy to purchase.

CONCLUSION Cr(VI) reduction by photocatalysis and photochemistry processes were studied, as well as their application in wastewater remediation. Photochemistry with ethanol under UV radiation was more efﬁcient than photochemistry with ethanol under visible light. Photocatalytic reduction with TiO2 in the presence of ethanol under UV radiation presented high Cr(VI) removals, showing the synergistic effect between ethanol oxidation and Cr(VI) reduction. Furthermore, results obtained with raw wastewater in its original concentration suggested high values of total Cr(VI) reduction, thus potentially conﬁrming the feasibility of applying these processes to remove hexavalent chromium from wastewater discharges of electroplating industries. Among the processes observed, the most suitable appears to be photochemistry with ethanol under UV radiation.

ACKNOWLEDGEMENTS The authors gratefully acknowledge the ﬁnancial support of CAPES and CNPq.

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2014

REFERENCES Chakrabarti, S., Chaudhuri, B., Bhattacharjee, S., Ray, A. K. & Dutta, B. K.  Photo-reduction of hexavalent chromium in aqueous solution in the presence of zinc oxide as semiconductor catalyst. Chemical Engineering Journal 153, 86–93. Cho, Y., Kyung, H. & Choi, W.  Visible light activity of TiO2 for the photoreduction of CCl4 and Cr(VI) in the presence of nonionic surfactant (Brij). Applied Catalysis B: Environmental 52, 23–32. Das, D. P., Parida, K. & De, B. R.  Photocatalytic reduction of hexavalent chromium in aqueous solution over titania pillared zirconium phosphate and titanium phosphate under solar radiation. Journal of Molecular Catalysis A: Chemical 245, 217–224. Jiang, F., Zheng, Z., Xu, Z., Zheng, S., Guo, Z. & Chen, L.  Aqueous Cr(VI) photo-reduction catalyzed by TiO2 and sulfated TiO2. Journal of Hazardous Materials B134, 94–103. Ku, Y. & Jung, I.-L.  Photocatalytic reduction of Cr(VI) in aqueous solutions by UV irradiation with the presence of titanium dioxide. Water Research 35, 135–142. Mohapatra, P., Samantaray, S. K. & Parida, K.  Photocatalytic reduction of hexavalent chromium in aqueous solution over sulphate modiﬁed titania. Journal of Photochemistry and Photobiology A: Chemistry 170, 189–194. Mytych, P. & Stasicka, Z.  Photochemical reduction of chromium(VI) by phenol and its halogen derivatives. Applied Catalysis B: Environmental 52, 167–172. Mytych, P., Karocki, A. & Stasicka, Z.  Mechanism of photochemical reduction of chromium(VI) by alcohols and its environmental aspects. Journal of Photochemistry and Photobiology A: Chemistry 160, 163–170. Nasrallah, N., Kebir, M., Koudri, Z. & Trari, M.  Photocatalytic reduction of Cr(VI) on the novel heterosystem CuFe2O4/CdS. Journal of Hazardous Materials 185, 1398–1404. Schrank, S. G., José, H. J. & Moreira, R. F. P. M.  Simultaneous photocatalytic Cr(VI) reduction and dye oxidation in a TiO2 slurry reactor. Journal of Photochemistry and Photobiology A: Chemistry 147, 71–76. Tuprakay, S. & Liengcharernsit, W.  Lifetime and regeneration of immobilized titania for photocatalytic removal of aqueous hexavalent chromium. Journal of Hazardous Materials B124, 53–58. Xu, X., Li, H. & Gu, J.  Simultaneous decontamination of hexavalent chromium and methyl tert-butyl ether by UV/ TiO2 process. Chemosphere 63, 254–260. Yang, G. C. C. & Chan, S.-W.  Photocatalytic reduction of chromium(VI) in aqueous solution using dye-sensitized nanoscale ZnO under visible light irradiation. Journal Nanoparticles Research 11, 221–230. Yurkow, E. J., Hong, J., Min, S. & Wang, S.  Photochemical reduction of hexavalent chromium in glycerol-containing solutions. Environmental Pollution 117, 1–3.

First received 10 December 2013; accepted in revised form 8 April 2014. Available online 22 April 2014

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Removal and retention of phosphorus by periphyton from wastewater with high organic load Jinxiang Cao, Xiaoxing Hong and Guofeng Pei

ABSTRACT The total phosphorus (TP) removal efficiency from organic wastewater (pig farm and distillery wastewater) were estimated by using filamentous green algae (FGA) and benthic algal mats (BAM) treatment systems under laboratory conditions, and the contents of periphyton phosphorus fractions were determined by using a sequential extraction. The removal rates of TP reached 59–78% within the first 8 days of all treatment systems and could achieve average 80% during 30 day period,

Jinxiang Cao Xiaoxing Hong Guofeng Pei (corresponding author) College of Life Sciences, South-Central University for Nationalities, Wuhan, China E-mail: peigf@mail.scuec.edu.cn

and the phosphorus removal rates by using BAM was higher than that of FGA. The ability of retention TP of periphyton enhanced gradually, the BAM TP contents were higher than that of FGA, the highest TP concentrations of BAM and FGA were 26.24 and 10.52 mg P g 1·dry weight. Inorganic phosphorus (IP) always exceeded 67.5% of TP, but the organic phosphorus fraction only made up less than 20% of TP. The calcium-binding phosphorus (Ca-P) was the dominant fraction and its relative contribution to TP was more than 40%. The TP was also strongly and positively correlated with the IP and Ca-P (p < 0.01) in periphyton. It showed that the periphyton had a potential ability of rapid phosphorus removing and remarkable retention from wastewater with high load phosphorus. Key words

| periphyton, phosphorus fraction, phosphorus retention, removal rate, wastewater treatment

INTRODUCTION Eutrophication due to excess phosphorus (P) is a persistent condition in surface waters and a widespread environmental problem (Carpenter ). Thus, it is necessary to reduce total phosphorus (TP) for eutrophic water column. Environmental researchers have proposed and applied many phosphorus removal methods, including algae-based wastewater treatment. This treatment method has been investigated for decades because of its effectiveness in reducing the possibility of eutrophication and other ecological damage (Silva-Benavides & Torzillo ). Many researchers used microalgae (mixed or monocultures) for wastewater purification in high-rate algal ponds (Guzzon et al. ), but challenges are involved in microalgae harvesting (Shi et al. ). Unlike microalgae, filamentous green algae (FGA) and benthic algal mats (BAM) can be controlled well and harvested easily. The major part of a floating wetland is a floating raft or mat (Chua et al. ). These mats had key functions in nutrient uptake, cycling, and soil/sediment formation (McCormick et al. ), and their biomass has been acknowledged as a potentially important shortdoi: 10.2166/wst.2014.195

term sink for phosphorus (McCormick et al. ). Therefore, some interesting and meaningful results could be found in the application of periphyton to wastewater treatment. Periphyton can remove phosphorus from water effectively and utilize phosphorus from sediments, which are important regulating factors for the exchange of phosphorus between sediment and overlying water (Wetzel ). Gaiser et al. () found that measurements of phosphorus in periphyton mats could be used as an early signal of enhanced phosphorus load. However, almost no information is available in the literature about periphyton phosphorus forms and their measuring method. The Standard Measurements and Testing (SMT) sequential extraction method was used to determine phosphorus content in sediment and to group sediment phosphorus into ﬁve forms (Ruban et al. ; Zhu et al. ): (i) TP, (ii) phosphorus as organic esters of phosphoric acid (OP), (iii) inorganic phosphorus (IP), (iv) apatite phosphorus in forms associated with Ca (Ca-P or HCl-P), and (v) non-apatite IP in forms associated

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Removal and retention of P by periphyton from wastewater

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2014

with oxides and hydroxides of Fe, Al, and Mn (Fe-P or NaOH-P). Algal assemblages can inﬂuence phosphorus cycling either directly (uptake and release) or indirectly through photosynthesis-induced changes in water and sediment/ water interface parameters (Vymazal ). Thus, in order to assess the risk of eutrophication in aquatic systems, determining the content and distribution of TP among the different sediment phases is necessary (Gonzalez Medeiros et al. ). In the present work, two kinds of periphyton treatment system (FGA and BAM) were designed to measure and compare the phosphorus removal efﬁciency from wastewater with high organic load. Meanwhile the concentration of phosphorus fractions in periphyton was determined by using the SMT protocol. The objective was to evaluate the ability of phosphorus removal for periphyton and the retained phosphorus forms of periphyton under different conditions.

greenhouse under continuous cool-white fluorescent light illumination of approximately 100–200 μmol photons m 2 s 1. The temperature was controlled at approximately 20 ± 2 C. All containers were continuously supplied with air at a flow rate of 6 L min 1 to provide sufficient CO2 for algal growth. The initial pH value of the cultures was controlled to neutral (7.0–7.5).

MATERIALS AND METHODS

Water and periphyton sampling

Periphyton and wastewater collection

Efﬂuent samples (100 mL) were taken at intervals of 4 days for the ﬁrst 20 days and at intervals of 10 days thereafter. These samples were analyzed immediately for TP and soluble reactive phosphorus (SRP) based on Standard Methods for the Examination of Water and Wastewater (APHA ). The pH value was measured in situ using hand-held probes. Periphyton samples (8 g wet weight) were taken at intervals of 10 days and analyzed for phosphorus fractions according to the SMT sequential extraction method of the European Commission (Ruban et al. ). The content of different phosphorus forms were reported per unit dry weight (DW). All analyses of phosphorus fractions were carried out in triplicate. The

FGA were obtained from a local freshwater lake. BAM obtained from artiﬁcial substrates (granite stone: 15 × 8 × 1 cm) were cultured for 20 days in the littoral zone (0.2– 0.5 m deep) of a shallow eutrophic lake. Pig farm wastewater and distillery wastewater were provided by a local pig farm and distillery, respectively. Culture conditions Plexiglass cylinders were used as experimental containers (40 × 26 × 28 cm3). Cultivations were carried out in a Table 1

Experimental design The primary concentration of FGA and BAM were about 10 and 5 g L 1 (wet weight), respectively, in each treatment system. Seven groups (A–G) were performed in this study (Table 1). At the beginning of each experiment, all periphyton were grown in BG-11 medium without phosphorus for 1 week. The periphyton culture solution comprising organic wastewater and distilled water, and the total volume was 15 L. The experiment lasted for 30 days.

Experimental design of organic wastewater treatment systems with periphyton Initial concentration of P (μg mL 1)

Experimental period (days)

Number

Algal culture

A (control)

Pig farm wastewater (PW)

8.31

B (control)

Distillery wastewater (DW)

7.73

C (control)

FGA þ tap water

FGA þ PW

5.65

FGA þ DW

10.24

BAM þ DW

10.82

FGA ¼ ﬁlamentous green algae; BAM ¼ benthic algal mats.

0.06

10.91

J. Cao et al.

Removal and retention of P by periphyton from wastewater

SRP concentrations were analyzed using the molybdenum blue method. Statistical analysis Data analyses were conducted using SPSS 17.0. Differences in phosphorus fraction of the periphyton samples were tested using repeated-measures analysis of variance (ANOVA). All data were transformed by the natural logarithm to improve normality and homogeneity of variance before analysis. The data of phosphorus fraction were tested for correlation with other phosphorus properties using Pearson correlation coefﬁcients. The phosphorus removal rate was calculated using the equation Rn ¼ (C0 Cn)/C0*100%, where C0 and Cn are the concentrations of P at the beginning and the nth day of the treatment, respectively.

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2014

Figure 1, the pH value of the water column varied from 7.1 to 9.2. In general, the pH value steadily increased and pertained to slight alkalinity during the experimental period. No signiﬁcant changes in pH value were noted between the pig farm wastewater and distillery wastewater (p < 0.01), and the FGA and BAM system was also similar (p < 0.01). At the end of the experiment, the pH value of all systems associated with the photosynthesis of periphyton was higher than 8.5. Removal of phosphorus from organic wastewater The TP and SRP removal by FGA and BAM are shown in Figure 2. The TP and SRP concentration of organic wastewater were signiﬁcantly decreased, especially in the initial 8 days. However, the concentration of phosphorus in the control slightly decreased. At the end of the experiment, the TP

RESULTS Growth of periphyton Most ﬁlamentous algae were found ﬂoating within the upper 20 cm of the water column. FGA were mainly composed of Pithophora oedogonia, and the dominant phyla of BAM included Cyanophyta (Oscillatoria spp. and Phormidium spp.), Chlorophyta (Stigeoclonium spp.), and Bacillariophyta (Navicula spp. and Gomphonema spp.). Organic wastewater was highly turbid at the beginning of culture, and became clearer after 8 days of growth. As shown in

Figure 2 Figure 1

Changes in the pH value of wastewater (groups A–G) during the culture period.

Concentrations of water column TP and SRP of experimental groups (A–G) during a 30 day period.

J. Cao et al.

Removal and retention of P by periphyton from wastewater

concentrations of the pig farm wastewater-treated groups (D and E) reduced to the same level, ranging from 10.97 to 1.94 μg mL 1 and from 5.6 to 1.76 μg mL 1, respectively. In the distillery wastewater-treated groups (F and G), the TP concentrations were decreased from 10.25 to 3.06 μg mL 1 by FGA and from 10.79 to 1.67 μg mL 1 by BAM, respectively. During the experiment, changes of TP and SRP concentrations showed satisfactory phosphorus removal rates (Figure 3). The TP removal rate reached 59–78% within the ﬁrst 8 days of adding wastewater for FGA and BAM, respectively, and the highest value (84.7%) was found in group G. The FGA with high concentrations of pig farm wastewater had a higher removal rate (group D, 82.2%) than that of low concentrations group E (70.0%). For the different organic wastewater, groups E and F (70.2%) showed similar TP removal rates. The SRP removal rates were strongly correlated with the TP removal rates (r ¼ 0.989, p ¼ 0.011). Phosphorus fractions in periphyton The periphyton phosphorus fractions contents were analyzed (Figure 4). Compared with control group C, the contents of TP, IP, and Ca-P steadily increased in the 30 day culture period, whereas the concentrations of OP and Fe-P had peak values after 20 days and decreased thereafter. The BAM exhibited remarkable phosphorus retention capacity, with TP contents ranging from 3.51 ± 0.11 to 26.24 ± 1.10 mg P g 1 DW, it achieved the highest TP content compared to the FGA (2.29 ± 0.12 to 10.52 ± 0.59 mg P g 1 DW).

Figure 3

The removal rates of TP by periphyton treatment groups from organic wastewater.

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2014

The contents of OP and Fe-P were low, accounting only for a small portion of TP. Signiﬁcant differences for phosphorus fractions were noted among the experimental groups via repeated-measures ANOVA (p < 0.01). The phosphorus forms in FGA and BAM were mainly associated with inorganic forms. The sum of phosphorus contents in the IP and OP fractions agreed (94.7–99.6%, groups D–G) with the TP content. The IP was always greater than 67.5% of TP; thus, it was considered to be the main fraction of TP in all periphyton samples. Ca-P (40–54% of TP) was the dominant fraction in IP. The Fe-P and OP contents were 24–34% and 12–20% of the TP content, respectively (Figure 4). Pearson correlations (Table 2) showed that Ca-P and IP were strongly correlated with TP (r ¼ 0.982, 0.983; p < 0.01). Fe-P was only signiﬁcantly correlated with OP (r ¼ 0.618, p < 0.01), and no strong correlations were observed with other phosphorus fractions. The periphyton phosphorus fractions had a strongly negative correlation with the TP and SRP of water column (Table 2). The P contents of wastewater declined, while that of periphyton steadily increased for the period of culture.

DISCUSSION The phosphorus removal rate is correlated with initial phosphorus content (Pietro et al. ). Based on the results of the treatment groups, the group with higher initial P content also had greater phosphorus removal rate. Similar ﬁndings were reported in other studies (Pietro et al. ), where Ceratophyllum/periphyton complex SRP increased in a linear trend as initial phosphorus concentrations were increased (0–11 μg P mL 1) during incubation. Although ambient phosphorus contents at the initial period of each treatment changed, these levels decreased to almost the same level at the end of the experiment. The results of this study showed that phosphorus removal rates were only correlated with initial phosphorus concentration at the initial culture period. And the long-term phosphorus removal capacity of periphyton was related to algal growth environment, such as ambient phosphorus contents. Shi et al. () found that the differences of phosphorus removal depend on the type of wastewater, the type of periphyton, and their growth conditions. This relationship is related to the experimental process system, which was a relatively closed system in present study. The phosphorus in periphyton, phosphorus adsorption by algae, and ambient phosphorus achieved a dynamic equilibrium when the algae were cultured for a

J. Cao et al.

Figure 4

Table 2

Removal and retention of P by periphyton from wastewater

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Changes in periphyton phosphorus fraction concentrations of treatment groups (C–G) during the study period.

Pearson correlations between the periphyton phosphorus fractions and wastewater TP and SRP

H2O TP

H2O SRP

0.992**

H2O SRP

Ca-P

Fe-P

0.599**

0.636**

0.475*

0.488*

0.787**

0.609**

0.644**

0.983**

0.724**

Ca-P

0.542*

0.578**

0.982**

0.733**

0.963**

Fe-P

0.470*

–

0.618**

–

H2O TP ¼ wastewater TP (P); H2O SRP ¼ wastewater soluble reactive P; only signiﬁcant values are provided. * p < 0.05, ** p < 0.01.

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Removal and retention of P by periphyton from wastewater

specific period. In this study, it was found that the phosphorus content of wastewater rapidly declined at the initial period (0–8 days). The fact that both FGA and BAM can rapidly remove P from organic wastewater showed that these two treatment systems have important functions in short-term nutrient removal. Similar results were found in the Everglades (McCormick et al. ). Dodds () indicated that algal species with a high surface-area to volume ratio, such as filamentous algae, have greater nutrient uptake rates than prostrate gelatinous forms. However, compared to the suspension of FGA in all treatment groups, BAM exhibited improved P removal efficacy. This effect was the result of species composition differences. FGA were mainly composed of Chlorophyta species (e.g., P. oedogonia), whereas BAM had higher biodiversity, with Chlorophyta, Cyanophyta and Bacillariophyta species. Cardinale () found that communities with more species take greater advantage of the niche opportunities in an environment and diverse systems can capture a greater proportion of biologically available resources. Therefore, it was believed that the nutrient uptake capacity of periphyton is less related to the surface-area to volume ratio and more strongly to biodiversity. However, phosphorus removal was not completely attributed to periphyton because bacteria can utilize dissolved OP and IP (Scinto & Reddy ). The phosphorus content of the control groups (A and B) without periphyton slightly reduced, suggesting that the majority of phosphorus removal in treatments was associated with periphyton. The decrease in phosphorus in this study can explain the differences between the control and treatment cultures. The periphyton was involved in the transformation of water column phosphorus. They created the appropriate biochemical conditions for phosphorus removal or precipitation and quickly responded to replenish wastewater and to transfer P from wastewater again. The periphyton was considered by many researchers as a short-term sink for P (McCormick et al. ; Pei et al. ). The periphyton TP contents, which ranged from 1.54 to 26.24 mg g 1 DW, were higher than the values of surface sediment TP from two lakes of northern Greece, with 0.9–1.3 mg g 1 DW (Fytianos & Kotzakioti ), and higher than the values of periphyton TP from interior Everglades, with 0.104–0.289 mg g 1 DW (Scinto & Reddy ). The results in this study showed that both FGA and BAM had important functions in short-term nutrient storage. The storage capacity of phosphorus for periphyton was the equivalent of a buffer of an aquatic system after increased phosphorus loading. This buffering capability can rapidly decrease the concentration of phosphorus in wastewater

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and protect aquatic ecosystems against the ecological effects of excess phosphorus. The periphyton phosphorus forms were composed of two parts, including biotic, and abiotic uptake of phosphorus. The results displayed that the disappearance of phosphorus in the wastewater mainly attributed to the increasing of Ca-P induced by the elevation in pH value during periphyton incubation. Several authors reported that the increase in pH value (>8.5) was due to the fact that photosynthesis promotes the precipitation of phosphorus as insoluble salts (Silva-Benavides & Torzillo ). DeBusk et al. () found that most of the phosphorus removed by periphyton after 8 months of operation stored in new periphyton biomass. The present experiments only lasted for 1 month, and phosphorus storage in periphyton should principally result from abiotic uptake. The OP and Fe-P increased in the initial 20 days and then declined thereafter. The relative contribution of OP to total sedimentary P was also minor compared to Fe-P, the OP in the surﬁcial sediment decreased only slightly, suggesting a slow mineralization process during the entire sampling period (Amirbahman et al. ). Although direct information on the turnover of phosphorus from algal cells and the mechanisms responsible for OP and Fe-P depletion was lacking, these changes may be due to periphyton growth and decomposition by the end of the measurement period. Richardson et al. () assumed that algal biomass had a minor function in long-term phosphorus retention because of high turnover rates. During cell lysis and decomposition, most algal phosphorus was organic and underwent bacterial degradation (Wetzel ). In addition, some researchers proposed that OP can convert to IP by extracellular enzyme catalysis (DeBusk et al. ). Jin et al. () found that elevated pH value resulted in the release of Fe-P in lake sediments. The Fe-P reduction was possibly caused by elevated pH value. This result was beneﬁcial to phosphorus removal because the released cations and their hydroxides effectively served as coagulants for P removal (Seida & Nakano ). The results of this study indicated that periphyton reduced the residence time of excess phosphorus in water column and exhibited strong capability and potential for short-term phosphorus retention. Mulholland et al. () demonstrated that the accumulation of periphyton biomass could increase internal nutrient cycling within the periphyton matrix. Gaiser et al. () showed that phosphorus measurements in periphyton mats could be used as an early signal of enhanced phosphorus load and a criterion to measure ecosystem degradation. However, environmental parameters, such as pH value, incubation

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Removal and retention of P by periphyton from wastewater

time, temperature, and excessive aeration, had major functions in the removal of phosphorus from wastewater (deBashan & Bashan ). Moreover, more complex changes in trophic structure and hydraulics occurred in natural ecosystems. To understand the function of periphyton in phosphorus cycling under natural conditions, the role of periphyton in the phosphorus shift between overlying water and sediment should be investigated in eutrophic shallow lake.

CONCLUSIONS 1. The TP removal rates of wastewater with high organic load could reach 80% using periphyton treatment system, and the TP removal efﬁciency of BAM was higher than that of FGA. 2. The Ca-P fractions constituted more than 40% of periphyton TP. This maybe plays an important role of periphyton-mediated processes for the P-ﬂuxes among different P forms and the P-transformation. 3. The periphyton reduced the residence time of excess phosphorus in water column and exhibited strong capability and potential for short-term phosphorus retention. 4. Cultivating periphyton in organic wastewater presents an alternative to restore the eutrophic shallow water column, and will give some suggestions in managing phosphorus pollution of Chinese water column. In the future, it was hoped to perform a long-term and largescale experiment to study this problem in a suitable lake.

ACKNOWLEDGEMENTS This research was supported by the National Science Foundation of China (grant no. 30970550). We also thank Hai-bo Wang for his helpful assistance.

REFERENCES Amirbahman, A., Lake, B. A. & Norton, S. A.  Seasonal phosphorus dynamics in the surﬁcial sediment of two shallow temperate lakes: a solid-phase and pore-water study. Hydrobiologia 701 (1), 65–77. APHA  Standard Methods for the Examination of Water and Wastewater. American Public Health Association, Washington, DC, USA. Cardinale, B. J.  Biodiversity improves water quality through niche partitioning. Nature 472 (7341), 86–89.

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Carpenter, S. R.  Eutrophication of aquatic ecosystems: bistability and soil phosphorus. Proceedings of the National Academy of Sciences of the United States of America 102 (29), 10002–10005. Chua, L. H., Tan, S. B., Sim, C. H. & Goyal, M. K.  Treatment of baseflow from an urban catchment by a floating wetland system. Ecological Engineering 49, 170–180. de-Bashan, L. E. & Bashan, Y.  Recent advances in removing phosphorus from wastewater and its future use as fertilizer (1997–2003). Water Research 38 (19), 4222–4246. DeBusk, T. A., Grace, K. A., Dierberg, F. E., Jackson, S. D., Chimney, M. J. & Gu, B.  An investigation of the limits of phosphorus removal in wetlands: a mesocosm study of a shallow periphyton-dominated treatment system. Ecological Engineering 23 (1), 1–14. Dodds, W. K.  The role of periphyton in phosphorus retention in shallow freshwater aquatic systems. Journal of Phycology 39, 840–849. Fytianos, K. & Kotzakioti, A.  Sequential fractionation of phosphorus in lake sediments of northern Greece. Environmental Monitoring and Assessment 100 (1), 191–200. Gaiser, E. E., Scinto, L. J., Richards, J. H., Jayachandran, K., Childers, D. L., Trexler, J. C. & Jones, R. D.  Phosphorus in periphyton mats provides the best metric for detecting lowlevel P enrichment in an oligotrophic wetland. Water Research 38 (3), 507–516. González Medeiros, J. J., Pérez Cid, B. & Fernández Gómez, E.  Analytical phosphorus fractionation in sewage sludge and sediment samples. Analytical and Bioanalytical Chemistry 381 (4), 873–878. Guzzon, A., Bohn, A., Diociaiuti, M. & Albertano, P.  Cultured phototrophic biofilms for phosphorus removal in wastewater treatment. Water Research 42 (16), 4357–4367. Jin, X.-c., Wang, S.-r. & Pang, Y.  The influence of phosphorus forms and pH on release of phosphorus from sediments in Taihu Lake (in Chinese). China Environmental Science 24 (6), 707–711. McCormick, P. V. & O’Dell III, M. B. R.B.E.S., Backus, J. G. & Kennedy, W. C.  Periphyton responses to experimental phosphorus enrichment in a subtropical wetland. Aquatic Botany 71, 119–139. McCormick, P. V., Shuford, R. B. E. & Chimney, M. J.  Periphyton as a potential phosphorus sink in the Everglades Nutrient Removal Project. Ecological Engineering 27 (4), 279–289. Mulholland, P., Steinman, A., Marzolf, E., Hart, D. & DeAngelis, D.  Effect of periphyton biomass on hydraulic characteristics and nutrient cycling in streams. Oecologia 98 (1), 40–47. Pei, G.-f., Liu, G.-x. & Hu, Z.-y.  The role of benthic algae in phosphorus retention in Donghu Lake, Wuhan (in Chinese). Acta Scientiae Circumstantiae 29 (4), 840–845. Pietro, K. C., Chimney, M. J. & Steinman, A. D.  Phosphorus removal by the Ceratophyllum/periphyton complex in a south Florida (USA) freshwater marsh. Ecological Engineering 27 (4), 290–300.

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Richardson, C., Qian, S., Craft, C. & Qualls, R.  Predictive models for phosphorus retention in wetlands. Wetlands Ecology and Management 4 (3), 159–175. Ruban, V., López-Sánchez, J., Pardo, P., Rauret, G., Muntau, H. & Quevauviller, P.  Selection and evaluation of sequential extraction procedures for the determination of phosphorus forms in lake sediment. Journal of Environmental Monitoring 1 (1), 51–56. Ruban, V., Lopez-Sanchez, J., Pardo, P., Rauret, G., Muntau, H. & Quevauviller, P.  Harmonized protocol and certiﬁed reference material for the determination of extractable contents of phosphorus in freshwater sediments – A synthesis of recent works. Fresenius’ Journal of Analytical Chemistry 370 (2), 224–228. Scinto, L. J. & Reddy, K. R.  Biotic and abiotic uptake of phosphorus by periphyton in a subtropical freshwater wetland. Aquatic Botany 77 (3), 203–222. Seida, Y. & Nakano, Y.  Removal of phosphate by layered double hydroxides containing iron. Water Research 36 (5), 1306–1312.

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Shi, J., Podola, B. & Melkonian, M.  Removal of nitrogen and phosphorus from wastewater using microalgae immobilized on twin layers: an experimental study. Journal of Applied Phycology 19 (5), 417–423. Silva-Benavides, A. M. & Torzillo, G.  Nitrogen and phosphorus removal through laboratory batch cultures of microalga Chlorella vulgaris and cyanobacterium Planktothrix isothrix grown as monoalgal and as co-cultures. Journal of Applied Phycology 24 (2), 267–276. Vymazal, J.  Removal of nutrients in various types of constructed wetlands. Science of the Total Environment 380 (1), 48–65. Wetzel, R.  Limnology, 3 E. Lake and River Ecosystems. Academic Press, 525 B Street, Ste. 1900, San Diego, CA 92101–4495, USA. pp. 251–266. Zhu, M.-y., Zhu, G.-w., Li, W., Zhang, Y.-l., Zhao, L.-l. & Gu, Z.  Estimation of the algal-available phosphorus pool in sediments of a large, shallow eutrophic lake (Taihu, China) using proﬁled SMT fractional analysis. Environmental Pollution 173, 216–223.

First received 12 January 2014; accepted in revised form 8 April 2014. Available online 22 April 2014

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A pilot scale comparison of advanced oxidation processes for estrogenic hormone removal from municipal wastewater efﬂuent Radka Pešoutová, Luboš Strí̌ teský and Petr Hlavínek

ABSTRACT This study investigates the oxidation of selected endocrine disrupting compounds (estrone, 17βestradiol, estriol and 17α-ethinylestradiol) during ozonation and advanced oxidation of biologically treated municipal wastewater efﬂuents in a pilot scale. Selected estrogenic substances were spiked in the treated wastewater at levels ranging from 1.65 to 3.59 μg · L 1. All estrogens were removed by ozonation by more than 99% at ozone doses 1.8 mg · L 1. At a dose of 4.4 · mg L 1 ozonation reduced concentrations of estrone, 17β-estradiol, estriol and 17α-ethinylestradiol by 99.8, 99.7, 99.9 and 99.7%, respectively. All tested advanced oxidation processes (AOPs) achieved high removal rates but they were slightly lower compared to ozonation. The lower removal rates for all tested advanced

Radka Pešoutová (corresponding author) Luboš Strí̌ teský Petr Hlavinek AQUA PROCON s.r.o., Palackého tr.̌ 12, Brno, Czech Republic E-mail: radka.pesoutova@aquaprocon.cz Petr Hlavínek Faculty of Civil Engineering, Brno University of Technology, Brno, Czech Republic

oxidation processes are caused by the presence of naturally occurring hydroxyl radical scavengers – carbonates and bicarbonates. Key words

| advanced oxidation processes, AOPs, estrogenic hormones, ozone, wastewater treatment

INTRODUCTION Wastewater treatment plants (WWTPs) have always been one of the major sources of pollution for the aquatic ecosystems. Conventional activated sludge treatment technologies ensure high efficiency of biologically degradable organic pollution and nutrient removal, but are inefficient towards the so-called emerging contaminants: substances such as pharmaceuticals, pesticides, heavy metals, surfactants, burning inhibitors or cosmetic products (Crisp et al. ). Prior research proved that concentrations of estrogenic hormones in municipal effluents and in surface water are sufficient to initiate changes in endocrine system resulting in metabolic stress, risk of liver damage, etc. (Purdom et al. ; Arcand Hoy et al. ; Kujalová et al. ). Concentrations of the estrogenic hormones in surface water are in order of units to tens of ng · L 1 and major share falls to 17α-ethinylestradiol (Kuch & Ballschmiter ; Cargouet et al. ). As low as 0.1 ng · L 1 of 17αethinylestradiol is sufficient to alter the endocrine system of fish (Purdom et al. ). Despite the demonstrated negative impacts on aquatic fauna, no limits for estrogenic hormones concentrations in the WWTP effluents have been set by legislation to this doi: 10.2166/wst.2014.196

date. However, the European Union has been encouraging projects focusing on the research on emerging contaminants with the aim of comprehending their impacts on living organisms and to restrict their emissions into the environment. Tertiary stage of wastewater treatment based on advanced oxidation processes (AOPs) is a promising solution for removal of estrogenic hormones from municipal efﬂuents. A number of studies demonstrated effects of AOPs on degradation of emerging contaminants including estrogenic hormones (Ternes et al. ; Huber et al. ; Zhang et al. ; Larcher et al. ). The bulk of the previous studies address AOP trials in laboratory conditions. Scale-up and further implementation of emerging contaminants removal to common practice requires trials at pilot scale and on-site comparison of potential AOP technologies. This study compares ozonation and various AOPs (O3/H2O2; O3/UV; O3/UV/H2O2) as a tertiary step of treatment of municipal wastewater in a pilot scale on degradation of selected estrogenic hormones. Novelty of this work is the application of ozonation and different AOPs on a pilot scale closely to the real conditions in terms of the treated water quality.

R. Pešoutová et al.

Comparison of AOPs for estrogen removal from municipal efﬂuent

METHODS Pilot scale trials of ozonation and other AOP combinations took place at the WWTP Mikulov (municipality of Mikulov, Czech Republic). WWTP Mikulov is a conventional mechanical-biological treatment plant with capacity 24,850 population equivalents. The batch mode pilot plant (Figure 1) installed at the WWTP site was fed by treated wastewater from the secondary clarifier output. Treated effluent parameters were: pH ¼ 7.4 ± 0.1, chemical oxygen demand (CODCr) ¼ 30 ± 2.3 mg · L 1, biochemical oxygen demand (BOD5) ¼ 2.5 ± 1.1 mg · L 1, suspended solids ¼ 11 ± 2.2 mg · L 1, dissolved solids ¼ 970 ± 26 mg · L 1, Ntot ¼ 8.0 ± 5.1 mg · L 1; Ptot ¼ 1.1 ± 0.5 mg · L 1; absorbance436nm ¼ 0.017 ± 0.003; Escherichia coli ¼ 3800 ± 1900 CFU · 100 mL 1. Ozone generator EFFIZON GSO 10 (OZOMATIC) produced ozone from pure compressed oxygen (LINDE GAS) with the actual ozone production rate of 3.15 g O3 · h 1 and concentration of 5% of ozone in the gas mixture. A low pressure mercury lamp with an input of 230 W supplied UV radiation (254 nm) for the UV-based AOP trials. Membrane dosing pump (GRUNDFOS DDI 2–16) fed hydrogen peroxide of technical quality with concentration of 35% (supplied by VIA-REK) into the static mixer. A centrifugal pump (LOWARA, flow rate 4.3 m3 · h 1 at the given conditions) circulated the wastewater within the closed loop during each trial (batch) resulting in gradual increase of ozone (H2O2 and UV) dose in reaction tank with time. During trials the pilot plant discharge served as a point for taking samples. Based on previous studies the ozone dose of approximately a thousand-fold the sum of initial concentrations of estrogenic hormones were set to ensure the estrogenic

Figure 1

Scheme of the pilot plant.

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hormone removal rate is at least 95% at the end of each trial (Irmak et al. ; Zhang et al. ). The ozone transfer rate at given ozone generator output and wastewater volume was 0.088 mg O3 · L 1 · min 1 and it remained the same for all tested AOPs. The pH value of wastewater in all of the trials was 7.4 ± 0.1 and no further pH adjustments were made during trials. The O3:H2O2 weight ratio of 1:1.6 used in the O3/H2O2 and O3/UV/H2O2 trials is in the range suitable according to previous studies, although it is lower than stoichiometric ratio for peroxone process (USEPA ; Kurbus et al. ; Acar & Ozbelge ). Before each trial the wastewater was spiked with concentrated solution of selected estrogenic hormones (17βestradiol, estrone, estriol and 17α-ethinylestradiol) prepared by solving in DMSO (purity >99.9%). All of these were supplied by SIGMA-ALDRICH in a solid crystalline or powder form with purities: estrone >99%; estriol >97%; 17βestradiol >97%; and 17α-ethinylestradiol >98%. Analytical determinations of estrogenic hormones using liquid chromatography-tandem mass spectrometry (LC/MS/ MS) were performed at the RECETOX centre of the Masaryk University. Analytical determinations were performed as described by Cˇ echová & Šimek () with formic acid concentration of 5 mM and additional cleaning of eluate using Florisil Strata FL-PR (500 mg, 3 cm3) and Sep-Pak Vac NH2, (500 mg, 6 cm3, 55–105 μm). Limits of quantiﬁcation for estrone, 17βestradiol, 17α-ethinylestradiol and estriol were 1.4 ng · l 1, 1.0 ng · l 1, 1.1 ng · l 1 and 0.7 ng · l 1, respectively. All numerical values shown in this paper are mean values of the two duplicates. The differences between these two duplicates for all trials were lower than the differences in performances of the tested AOPs.

R. Pešoutová et al.

Comparison of AOPs for estrogen removal from municipal efﬂuent

RESULTS AND DISCUSSION Initial hormone concentrations and applied doses of chemicals The concentrations of selected estrogenic hormones in the municipal effluent were in order of ng · L 1. When considering typical limits of quantification of available commercial laboratories it was decided to spike municipal effluent with a mix of concentrated estrogenic hormones. The approximate sum of their concentrations at the beginning of trials was set to be approximately 8 μg · l 1. The selected estrogenic hormones were spiked to the wastewater to be able to determine 95 to 99% removal. The initial concentrations of estrogenic hormones for testing of ozonation and for testing of selected combinations of AOPs are presented in the following table (Table 1).

Effect of ozonation on estrogenic hormone degradation The ozonation trials results demonstrated high degradation efficiency for all of the monitored estrogenic hormones even at low ozone doses. Estrone was the fastest reacting substance; its concentration was reduced by 95.8% at the dose of 0.88 mg O3 · L 1 although its removal in a conventional biological treatment plant is the least of the tested estrogenic hormones ( Joss et al. ). As an intermediate product of 17β-estradiol, 17β-estradiol, estriol and 17α-ethinylestradiol at the end of the ozonation trials (final dose of 4.4 mg O3 · L 1) reached 99.8%, 99.7%, 99.9% and 99.7%, respectively (Figure 2). The degradation rates acquired in this study are comparable to other studies treating similar wastewaters and hormones with similar initial concentrations (Irmak et al. ; Zhang et al. ). Studies on low initial concentration of estrogenic hormones in municipal effluents (units of ng · L 1) showed a great variation in removal efficiencies

Table 1

Initial concentration of estrogenic hormones in spiked wastewater for testing of ozonation and for testing of combinations of selected AOPs 17βEstrone 1

[μg · L

]

17α-

Estradiol 1

[μg · L

]

Ethinylestradiol

Estriol 1

[μg · L

]

[μg · L

1.65

2.01

1.97

2.54

O3/H2O2

1.79

1.72

2.06

2.15

O3/UV

2.13

2.64

2.10

2.97

O3/H2O2/UV

1.95

2.88

2.52

3.59

]

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by ozonation (Ternes et al. ; Nakada et al. ). The weight ratio of degraded estrogenic hormone to ozone dose is (level of degradation in brackets): 1:560 for estrone (95.8%), 1:2,700 for estrone (99.8%), 1:2,200 for 17β-estradiol (99.7%), 1:560 for estriol (80.3%) and 1:1,700 for 17α-ethinylestradiol (99.7%).

Effect of UV on estrogenic hormone degradation during O3 treatment The results of O3/UV trials demonstrated no positive effect of UV irradiation on the removal rate of selected estrogenic hormones at given conditions. The removal rate of the ozonation exceeded the removal rate of O3/UV for all tested estrogenic hormones at all ozone doses (Figure 2). The ozone dose of 0.88 mg O3 · L 1 and the UV dose of 0.07 Wh · L 1 resulted in removal rate between 79.9% for estrone and 58.1% for estriol. The weight ratio of degraded estrogenic hormone to ozone dose is (level of degradation in brackets; for UV dose see Figure 2): 1:2,100 for estrone (99.4%), 1:1,700 for 17β-estradiol (98.2%), 1:2,100 for estriol (98.7%) and 1:1,500 for 17α-ethinylestradiol (99.6%). Generally lower removal rates are inconsistent towards other studies (e.g. Irmak et al. ) and were caused by the presence of radical scavengers (carbonates and bicarbonates) common to real wastewaters as described by von Gunten ().

Effect of H2O2 on estrogenic hormone degradation during O3 treatment The peroxone process (O3/H2O2) resulted in removal rate signiﬁcantly lower than the ozonation itself (see Figure 2). One potential cause of inhibition of degradation process can be exceeding the steady-state concentration of hydrogen peroxide resulting in the absorption of hydroxyl radicals by H2O2 (von Gunten ). This was not the case because the actual ratio of O3:H2O2 (1:1.6) was lower than the stoichiometric ratio (1:2.8). Another potential cause is the presence of hydroxyl radical scavengers. The weight ratio of degraded estrogenic hormone to ozone dose is (level of degradation in brackets; for H2O2 dose see Figure 2): 1:2,500 for estrone (99.4%), 1:2,600 for 17β-estradiol (98.2%), 1:2,200 for estriol (98.7%) and 1:2,100 for 17α-ethinylestradiol (99.6%).

Figure 2

R. Pešoutová et al.

Comparison of AOPs for estrogen removal from municipal efﬂuent

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Comparison of ozonation and different AOP combinations for estrogenic hormone degradation in dependence on O3, H2O2 and UV doses: (a) estrone, (b) 17β-estradiol, (c) estriol, (d) 17α-ethinylestradiol.

R. Pešoutová et al.

Comparison of AOPs for estrogen removal from municipal efﬂuent

Effect of H2O2 and UV on estrogenic hormone degradation during O3 treatment In theory combination of O3/H2O2/UV supports spontaneous ozone decomposition, thus improves the removal rate (Ultrox International ). In comparison to ozonation, the removal rate of O3/H2O2/UV trials decreased. The removal rate using this AOP is the worst of all for 17α-ethinylestradiol and estrone. The total removal rate at the end of the trials was comparable to other processes (removal rate exceeded 99%). Similarly to the peroxone (O3/H2O2) AOP an excessive amount of H2O2 may result in process inhibition (Ultrox International ). The actual ratio of O3:H2O2 was the same as in the O3/H2O2 trials hence the probable reason for inhibition was the presence of hydroxyl radical scavengers (carbonates and bicarbonates) and not exceeding the steady-state ratio of O3:H2O2. The weight ratio of degraded estrogenic hormone to ozone dose is as follows (level of degradation in brackets; for H2O2 and UV dose see Figure 2): 1:2,300 for estrone (99.4%), 1:1,500 for 17β-estradiol (98.2%), 1:1,800 for estriol (98.7%) and 1:1,200 for 17α-ethinylestradiol (99.6%).

CONCLUSIONS The effect of ozonation and three AOPs tested on the removal of estrogenic hormones from biologically treated municipal wastewater on pilot scale was investigated in this study. The following conclusions were made from the data obtained:

• • • • • • •

Estrogenic hormones are generally easily removed by oxidation. In contrast to other studies, the most efﬁcient process was ozonation. Disagreement with other studies was caused by presence of hydroxyl radicals scavengers (in this case the carbonates and bicarbonates). The presence of scavengers in real wastewaters results in AOPs being ineffective for removal of pollutants easily oxidised by ozonation (i.e. estrogenic hormones). The least resistant estrogenic hormone in all trials was synthetic 17α-ethinylestradiol. Natural hormone estrone was the most resistant estrogenic hormone tested. The lower the estrogenic hormone concentration the higher the required ratio of ozone to estrogenic hormone. For example when reducing estrone from 1.65 μg · L 1 to

• •

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0.07 μg · L 1 (i.e. reduction by 96%) the needed ratio is 1:560. When reducing estrone from 0.07 μg · L 1 to 0.003 μg · L 1 (i.e. reduction by 96%) the needed ratio is 1:56,000. The conclusion above could make real-life estrogenic hormone removal by ozonation ozone ineffective and/or cost ineffective as shown in Nakada et al. () and Ternes et al. (). Further research on this topic with focus on treatment of real-life municipal effluents without spiking the wastewater is required in order to confirm or disprove if ozonation is sufficient and economic solution for removal of estrogenic hormones from municipal effluents.

ACKNOWLEDGEMENTS This study was performed as part of the project AOP4Water within the framework Era-Net/CORNET for transnational collective research. The ﬁnancial support provided by a grant of the Ministry of Industry and Trade of the Czech Republic via Operational Programme Industry and Enterprise (OPIE) is greatly acknowledged. The authors also acknowledge DISA, s.r.o. for provision of pilot plant components and assistance during the trials and Vodovody a kanalizace Brě clav, a.s. for the access to the study area.

REFERENCES Acar, A. & Ozbelge, T.  Oxidation of acid red-151 aqueous solutions by the peroxone process and its kinetic evaluation. Ozone: Science & Engineering 28 (3), 155–164. Arcand-Hoy, L. D., Nimrod, A. C. & Benson, W. H.  Endocrine-modulating substances in the environment: estrogenic effects of pharmaceutical products. International Journal of Toxicology 17, 139–158. Cargouet, M., Perdiz, D., Mouatassim-Souali, A., TamisierKarolak, S. & Levi, Y.  Assessment of river contamination by estrogenic compounds in Paris area (France). Science of the Total Environment 324, 55–66. Č echová, E. & Šimek, Z.  Možnosti kvantitativního stanovení estrogenů v povrchových a odpadních vodách technikou HPLC-MS-MS (Possibility of Quantitative Determination of Estrogen in Surface and Waste Water Technology HPLC-MSMS). Chemické Listy 106 (S1), 3–6. Crisp, T. M., Clegg, E. D., Cooper, R. L., Wood, W. P., Anderson, D. G., Baetcke, K. P., Hoffmann, J. L., Morrow, M. S., Rodier, D. J., Schaeffer, J. E., Touart, L. W., Zeeman, M. G. & Patel, Y. M.  Environmental endocrine disruption: an effects assessment and analysis. Environment Health Perspectives 106 (Supplement 1), 11–56.

R. Pešoutová et al.

Comparison of AOPs for estrogen removal from municipal efﬂuent

Huber, M., Gbel, A., Joss, A., Hermann, N., Löffler, D., McArdell, C. S., Riedl, A., Siegrist, H., Ternes, T. A. & von Gunten, U.  Oxidation of pharmaceuticals during ozonation of municipal wastewater effluents: a pilot study. Environmental Science & Technology 39 (11), 4290–4299. Irmak, S., Erbatur, O. & Akgerman, A.  Degradation of 17βestradiol and bisphenol A in aqueous medium by using ozone and ozone/UV techniques. Journal of Hazardous Materials 126 (11), 54–62. Joss, A., Keller, E., Alder, C. A., Göbel, A., McArdell, C. S., Ternes, T. & Siegrist, H.  Removal of pharmaceuticals and fragrances in biological wastewater treatment. Water Research 39 (14), 3139–3152. Kuch, H. M. & Ballschmiter, K.  Determination of endocrine-disrupting phenolic compounds and estrogens in surface and drinking water by HRGC-(NCI)-MS in the picogram per liter range. Environmental Science & Technology 35, 3201–3206. Kujalová, H., Sýkora, V. & Pitter, P.  Látky s estrogenním účinkem ve vodách (Compounds with estrogenic activity in waters). Chemické listy 101, 706–712. Kurbus, T., Majcen Le Marechal, A. & Brodnjak Vončina, D.  Comparison of H2O2/UV, H2O2/O3 and H2O2/Fe2þ processes for the decolorisation of vinylsulphone reactive dyes. Dyes and Pigments 58 (3), 245–252. Larcher, S., Delbès, G., Robaire, B. & Yargeau, V.  Degradation of 17α-ethinylestradiol by ozonation – Identification of the by-products and assessment of their

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estrogenicity and toxicity. Environment International 39 (1), 66–72. Nakada, N., Shinohara, H., Murata, A., Kiri, K., Managaki, S., Sato, N. & Takada, H.  Removal of selected pharmaceuticals and personal care products (PPCPs) and endocrine-disrupting chemicals (EDCs) during sand ﬁltration and ozonation at a municipal sewage treatment plant. Water Research 41 (19), 4373–4382. Purdom, C. E., Hardiman, P. A., Bye, V. J., Eno, N. C., Tyler, C. R. & Sumpter, J. P.  Estrogenic effects of efﬂuents from sewage treatment works. Chemistry and Ecology 8, 275–285. Ternes, T. A., Stüber, J., Hermann, N., McDowell, D., Riedl, A., Kampmann, M. & Teiser, B.  Ozonation: a tool for removal of pharmaceuticals, contrast media and musk fragrances from wastewater. Water Research 37 (8), 1976–1982. Ultrox International  Ultrox International Ultraviolet Radiation/Oxidation Technology, Application Analysis Report. United States Environmental Protection Agency. Available at: http://nepis.epa.gov/Exe/ZyPURL.cgi? Dockey=10001TAV.TXT (accessed on 27 February 2014). USEPA  Handbook on Advanced Nonphotochemical Oxidation Processes BiblioBazaar, Charleston. 136. von Gunten, U.  Ozonation of drinking water. Part I. Oxidation kinetics and product formation. Water Research 37, 1443–1467. Zhang, X., Chen, P., Wu, F., Deng, N., Liu, J. & Fang, T.  Degradation of 17α-ethinylestradiol in aqueous solution by ozonation. Journal of Hazardous Materials 133 (1–3), 291–298.

First received 15 January 2014; accepted in revised form 8 April 2014. Available online 21 April 2014

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Vertical ﬂow constructed wetlands: kinetics of nutrient and organic matter removal M. M. Pérez, J. M. Hernández, J. Bossens, T. Jiménez, E. Rosa and F. Tack

ABSTRACT The kinetics of organic matter and nutrient removal in a pilot vertical subsurface wetland with red 2

ferralitic soil as substrate were evaluated. The wetland (20 m ) was planted with Cyperus alternifolius. The domestic wastewater that was treated in the wetland had undergone a primary treatment consisting of a septic moat and a buffer tank. From the sixth week of operation, the performance of the wetland stabilized, and a significant reduction in pollutant concentration of the effluent wastewater was obtained. Also a significant increase of dissolved oxygen (5 mg/l) was obtained. The organic matter removal efficiency was greater than 85% and the nutrient removal efficiency was greater than 75% in the vertical subsurface wetland. Nitrogen and biochemical oxygen demand (BOD) removal could be described by a first-order model. The kinetic constants were 3.64 and 3.27 d 1 for BOD and for total nitrogen, respectively. Data on the removal of phosphorus were

M. M. Pérez (corresponding author) J. M. Hernández T. Jiménez E. Rosa Study Center of Applied Chemistry, Chemical Pharmacy Faculty, Central University of Las Villas, Cuba E-mail: mairapv@uclv.edu.cu J. Bossens F. Tack Department of Applied Analytical and Physical Chemistry, Faculty of Bioscience Engineering, Ghent University, Belgium

adapted to a second-order model. The kinetic constant was 0.96 (mg/l) 1 d 1. The results demonstrated the potential of vertical ﬂow constructed wetlands to clean treated domestic wastewater before discharge into the environment. Key words

| kinetic model, vertical subsurface wetlands, wastewater puriﬁcation

INTRODUCTION Constructed wetlands can be employed as natural methods for wastewater treatment. These systems offer effective and reliable treatment of wastewater in a simple and inexpensive way, requiring minimal maintenance. The operational and maintenance costs may be affordable, even in developing countries. Such methods have certain advantages over the conventional treatment systems: they can be established at the location where the wastewater is produced, they have low-energy requirements compared to conventional wastewater treatment and they can be maintained by relatively untrained personnel (Tietz et al. ; Llorens et al. ). The knowledge about complex processes that occur in the plants, microorganisms and soil matrix in the wetlands is still incomplete. There is a need for a better appraisal on the purifying processes that occur in wetlands under different climatic and operational conditions. In consideration of their good efﬁciency and relatively small land requirement, vertical ﬂow constructed wetlands have been increasingly used as a popular alternative for the treatment of a wide range of wastewaters (Brix & Arias ; Prochaska & Zouboulis ; Langergraber doi: 10.2166/wst.2014.183

et al. ), achieving removal of total suspended solids (TSS, 90%), chemical oxygen demand (COD, 90%) and ammonia (NH4-N, 90%) (Brix & Arias ; Goldberg & Sposito ). However, phosphate (PO4-P) removal is fairly low, typically 20–30% (Prochaska & Zouboulis ). This removal can increase with the use of substrates that favor the retention of this compound (Tang et al. ; Lee et al. ; Zhao et al. ). The treatment performance in these systems depends on a variety of operational factors related to the system itself, the properties of the water and the way it is applied to the bed, the substrate type and size, the maturity of bed and the climate (Brix & Arias ; Rosenquit et al. ). The limitations of traditional gravel substrate for phosphorus removal in wetlands systems fostered the exploration of alternative substrates rich in Ca or Fe minerals (Wood & McAtamney ; Drizo et al. ; Arias et al.  Comeau et al. ; Vymazal ; Yalcuk & Ugurlu ). The most frequently employed modeling equation for wetlands systems is the ﬁrst-order model (Reed et al. ; Kadlec & Knight ). This model describes removal

M. M. Pérez et al.

Vertical ﬂow constructed wetlands: kinetics

of pollutants with time as an exponential decay function. However, often this model does not fit the complex operating conditions in the removal of main pollutants in the subsurface flow wetlands. To overcome the inadequacy of first-order models, a number of mechanistic models have been developed recently (Langergraber & Simunek ; Mayo & Bigambo ; Wang et al. ). Nevertheless, because they involve numerous empirical parameters, extensive data are needed to calibrate the model. Thus efficient prediction of removal rates under different circumstances is hampered (Rousseau et al. ). Efficient performance of the predictive models predominantly depends on these factors that are believed to play a major role on pollutant removal performances in wetland systems: differences in design parameters, environmental conditions and weather conditions, among others. However, the first-order model remains useful as a design equation for pollutant removal in constructed wetlands. The transition from first- to zero-order biological degradation kinetics due to increased load can be represented by Monod model kinetics, providing a close interrelation on substrate availability and growth of biomass; it could be an alternative solution for developing more realistic predictive models for wetlands (Saeed & Sun ). In Cuba, experience with the use of natural systems for wastewater treatment is limited, despite the benefits of low cost and ease of operation, and considering that most of the existing treatment plants currently are in a highly degraded state. The present paper aims to assess the performance of a vertical subsurface flow wetlands system under tropical conditions. The wetland was equipped with iron-rich soil, which is commonly found in the region, as the substrate. Kinetic constants for the removal of main pollutants in this system were obtained.

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) were collected and planted by hand with a range of four plants per m2. Iron-rich soil was used for substrate. A layer of 0.60 m depth was placed over a gravel (10–15 mm) layer of 0.15 m of depth, which aimed to prevent clogging of the perforations in the discharge pipe. The substrate contained a high content of iron, which favors phosphorus removal and a sorption capacity of 0.96 g/kg at 25 C (Pérez et al. , ; Pérez ). Also perforated pipes, as inlet and discharges structures, and vertical pipes were placed to oxygenate the deep layers. Clay was used as sealing material to line the treatment bed. The ﬂow of wastewater, pumped at rate of 0.375 m3/h, was fed intermittently in two batches per week lasting 4 h on each occasion. More detail about the construction of the wetland is shown in a technical schematic (Figure 1). The pilot vertical subsurface wetland operated with a superﬁcial loading rate of 19 g BOD/m2d. W

Batch feed kinetic experiment The wetland was monitored over a 12-week period to determine the performance and verify if performance was stable. The same parameters as in the influent wastewater were determined on the effluent. After stabilization of the wetland, the experiment to determine the kinetics of BOD, total Kjeldahl nitrogen and total phosphorus removal was started. First, the influent wastewater (Cin) was analyzed, and 15 m3 of influent wastewater was discharge into the wetland, after 10 min the outlet was closed. Two liters of effluent wastewater (Cout) were taken after 1, 2, 4, 8, 12 and 16 h. The valve at the outlet was opened to collect the effluent in a polyethylene recipient. The experiments were carried out four times in different periods, including summer and winter with a temperature range between 24 and 30 C (January, April, July and October 2011). The concentration data of nutrient and organic matter were fitted to the different kinetic models. The integrated form of the rate expressions for zero, first-order and second-order are given by Equations (1)–(3), respectively (Metcalf & Eddy ): W

MATERIALS AND METHODS Experimental pilot vertical subsurface wetland The experimental pilot vertical subsurface wetland is situated next to a weld rails industry in the central region of Cuba (Pérez et al. ). This industry does not use water in the production process. Their wastewaters are solely domestic. There is a primary treatment with a septic moat and a buffer tank for the wastewater. The primary treatment does not satisfy the Cuban regulatory limits for discharge (ONN ). The area of the experimental wetland is about 20 m2. Young Cyperus alternifolius, (Cui et al. ; Yao et al.

Cin Cout ¼ kt

(1)

where k ¼ reaction rate constant (mg l 1 d 1); ln

Cout ¼ kt Cin

(2)

M. M. Pérez et al.

Figure 1

Vertical ﬂow constructed wetlands: kinetics

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Cross section and top view of the pilot vertical subsurface wetland.

where k ¼ reaction rate constant (d 1); kt ¼

1 1 Cout Cin

(3)

where k ¼ reaction rate constant (mg l 1 d 1). The Monod model is represented by Zhao et al. (): kt ¼ K ln

Cin þ (Cin Cout ) Cout

(4)

Kjeldahl distillation method. Ammonia nitrogen was analyzed by titration with HCl, after collecting the distilled ammonia into boric acid. Colorimetric measurement were used for nitrite and total phosphorus determination with a UV Genesys 10 colorimeter (Thermo Electron Corporation, USA), Nitrite was analyzed by the diazotization method and total phosphorus by vanadium molybdate method. For nitrate determination a selective nitrate electrode (WTW 106625) was used. Redox potential was determined by potentiometric determination. TSS were determined after evaporating the sample to dryness at 103–105 C (APHA ). W

where k ¼ reaction rate constant (mg l 1 d 1), and K ¼ damping factor (mg/l). Chemical analysis

RESULTS AND DISCUSSION

The domestic wastewater samples, for the pilot units were taken and analyzed at regular times over 1 year. Prior to the experiment, effluent wastewater was sampled until week 12 to identify when the wetland performance was stabilized. pH and electrical conductivity were measured in the field using a portable meter (Hanna HI9811-5, Hanna Instruments, Romania). The following properties of the wastewater were determined according to the procedures described in Standard Methods (APHA ). COD was analyzed by a closed reflux titrimetric method. The BOD was determined according to the 5-d BOD test. Total Kjeldahl nitrogen was assessed by using the macro-

Pollutant removal capacity High removal efficiencies were achieved for most pollutants (Table 1). The efficiency of phosphorus removal was high compared to that commonly reported for gravel substrate 20–30% (Prochaska & Zouboulis ). This may result from the use of iron-rich soil as substrate. Iron oxides constitute an efficient sorption medium for P (Goldberg & Sposito ). The effluent concentrations for all parameters complied with the Cuban regulation standards (ONN ). There was an increase of dissolved oxygen in the wastewater effluent

M. M. Pérez et al.

Table 1

Vertical ﬂow constructed wetlands: kinetics

Water Science & Technology

Summary of characteristics of influent and effluent wetland wastewater (means, standard deviations n ¼ 4 and removal efficiency η)

Regulatory Parameter

level

Inﬂuent

Efﬂuent

6.5–8.5

7.3 ± 0.4

7.04 ± 0.02

–

Electrical conductivity (μS/cm)

1,400

690 ± 106

677 ± 18

–

COD (mg/l)

98 ± 17

10 ± 4

BOD5 (mg/l)

42 ± 6

6±4

(%)

Dissolved oxygen (mg/l) –

1.1 ± 0.4

5.4 ± 0.2

–

Total phosphorus (mg/l) 2

7.1 ± 0.8

1.7 ± 0.2

Kjeldahl nitrogen (mg/l) 5

14.0± 3.4

2.4 ± 0.2

NH4-N (mg/l)

–

2.6 ± 1.8

1.5 ± 0.2

NO2 nitrogen (mg/l)

–

NO3 nitrogen (mg/l)

–

5.0 ± 2.7

8.0 ± 0.3

–

Total suspended solid (mg/l)

–

20 ± 4

0.4 ± 0.1

Figure 2

Table 2

0.03 ± 0.02 0.03 ± 0.001

Removal of BOD, total Kjeldahl nitrogen and total phosphorus with time.

–

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(5.5 mg l 1) compared to the influent wastewater (1.1 mg l 1). An increase in the dissolved oxygen in the effluent is an advantage of vertical wetland, which has previously been reported by several authors (Brix & Arias ; Molle et al. ; Saeed & Sun ). Batch feed kinetic experiment Figure 2 shows the decrease in concentration with time of the three studied pollutants in the batch feed kinetic experiment. Table 2 shows the fitted equation and the determination coefficient for the four models studied. Both the removal of BOD and nitrogen were best described by the first-order model, according to Reed et al. () and Kadlec & Knight () although there was also a good fit of these parameters to the saturation model. These results are according to reports that BOD and nitrogen removal are dominated by biological breakdown through microorganisms adhering to soil and plant roots. In the literature, biological degradation is believed to mostly follow firstorder kinetics (Metcalf & Eddy ). Phosphorus removal was best described by a second-order model. Several authors agree that phosphorus removal does not follow a first-order model (Kadlec & Knight ; Sikdar ). The main removal mechanisms for phosphorus are chemical adsorption and precipitation (Kadlec & Knight ). The kinetic constants obtained for BOD, nitrogen and phosphorus are given in Table 3. The range of first-order reaction constants obtained for BOD was 3.22–4.06 d 1, which is within the range of 0.3–6.11 d 1 reported by Kadlec & Knight (). The variability in constants reported is mainly determined by difference in wetland design and climate and weather conditions. The first-order reaction constant obtained for nitrogen removal (3.26–3.38 d 1) by contrast far exceeds the range reported by Kadlec & Knight () (0.006–0.75 d 1). This may be due to the high temperature prevailing in

Kinetic model and correlation coefﬁcient

Zero order Cout Cin ¼ kt First order ln

Cout ¼ kt Cin

Second order

1 1 ¼ kt þ Cout Cin

Saturation

1 k 1 Cout Cin ¼ þ tln½Cout =Cin K K t

BOD

Total Kjeldahl nitrogen

Total phosphorus

y ¼ 1.6849x þ 13.675 R 2 ¼ 0.953

y ¼ 0.5221x þ 6.2289 R 2 ¼ 0.975

y ¼ 0.3118x þ 1.479 R 2 ¼ 0.88

y ¼ 0.1515x 0.2163 R 2 ¼ 0.997

y ¼ 0.1362x 0.3298 R 2 ¼ 0.999

y ¼ 0.026x 0.1265 R 2 ¼ 0.905

y ¼ 0.0113x þ 0.0148 R 2 ¼ 0.973

y ¼ 0.024x þ 0.0661 R 2 ¼ 0.978

y ¼ 0.045x þ 0.10 R 2 ¼ 0.989

y ¼ 0.0209x þ 0.0671 R 2 ¼ 0.978

y ¼ 0.0687x þ 0.0516 R 2 ¼ 0.996

y ¼ 0.111x þ 0.0511 R 2 ¼ 0.821

Table 3

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Vertical ﬂow constructed wetlands: kinetics

Total Kjeldahl nitrogen 1

k ¼ 3.64 ± 0.42 d

k ¼ 3.27 ± 0.11 d

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ACKNOWLEDGEMENT

Kinetic constant (24–30 C)

BOD

Water Science & Technology

Total phosphorus

k ¼ 1.08 ± 0.18 (mg/l) 1 d 1

this area, which are associated with an intensive microbial activity. An additional factor is that the flow in the vertical wetlands was intermittent, causing an alteration between aerobic and anaerobic conditions. The redox potential change in the vertical wetland bed, obtaining values around 650 mV in the most superficial areas and near to the roots of the plants and of around 100 mV in the deepest areas in the channel, favored the nitrification– denitrification process, which is the main removal mechanism proposed for nitrogen. With respect to the phosphorus removal, there is a general criterion in the literature on the variability of models and constants obtained (Kadlec & Knight ; Sikdar ), which are determined by the design of each particular system. It is mainly influenced by the type of substrate chosen. The range of second-order constants determined for phosphorus removal in the current wetland (0.90–1.26 (mg/l) 1 d 1) is also greater than generally reported elsewhere. This is probably due to high contents of iron minerals in the substrate used in the pilot wetland. Sikdar () reported values ranging between 0.031 and 0.31 (mg/l) 1 d 1 for horizontal systems and gravel as substrate.

CONCLUSIONS The vertical pilot wetland effectively decreased the concentrations of organic matter and nutrients in effluent wastewater. Removal efficiencies greater than 75% were achieved. A significant increase of the dissolved oxygen in the effluent wastewater was obtained. Kinetics of BOD and nitrogen removal were best described by a first-order model, which suggests that the main removal mechanism for this pollutant is biological. Phosphorus removal in the wetland was fitted better by a second-order model. Overall, the results demonstrated the potential of vertical flow wetlands to clean domestic effluents prior to their emission into the environment in tropical conditions.

The authors would like to thank the Cooperation Inter Universities Project, (VLIR), for supporting this research.

REFERENCES APHA, AWWA & WEF  Standard Methods for the Examination of Water and Wastewater, 21st edn, Washington, DC. Arias, C., del Bubba, M. & Brix, H.  Phosphorus removal by sands for use as media in subsurface flow constructed reed beds. Water Research 35 (5), 1159–1168. Brix, H. & Arias, C. A.  The use of vertical flow constructed wetlands for on-site treatment of domestic wastewater: New Danish guidelines. Ecological Engineering 25, 491–500. Comeau, Y., Brisson, J., Réville, J. P., Forget, C. & Drizo, A.  Phosphorus removal from trout farm effluents by constructed wetlands. Water Science and Technology 44, 55–60. Cui, L. H., Zhu, X. Z., Ouyang, Y., Chen, Y. & Yang, F. L.  Total phosphorus removal from domestic wastewater with Cyperus alternifolius in vertical-flow constructed wetlands at the microcosm level. International Journal of Phytoremediation 13 (7), 692–701. Drizo, A., Frost, C. A., Smith, K. A. & Grace, J.  Phosphate and ammonium removal by constructed wetlands with horizontal subsurface flow, using shale as a substrate. Water Science and Technology 35, 95–102. Goldberg, S. & Sposito, G.  A chemical model of phosphate adsorption by soils: I. Reference Oxide Minerals. Soil Science Society of American Journal 48, 772–778. Kadlec, R. H. & Knight, L.  Treatment Wetlands. Lewis Publishers, Boca Raton, FL. Langergraber, G. & Simunek, J.  Modeling variably saturated water flow and multi component reactive transport in constructed wetlands. Vadoze Zone Journal 4 (4), 924–938. Langergraber, G., Prandtstetten, C. H., Pressl, A., Rohrhofer, R. & Haberl, R.  Optimization of subsurface vertical flow constructed wetlands for wastewater treatment. Water Science and Technology 55 (7), 71–78. Lee, M. S., Drizo, A., Rizzo, D. M., Druschel, G., Hayden, N. & Twohig, E.  Evaluating the efficiency and temporal variation of pilot-scale constructed wetlands and steel slag phosphorus removing filters for treating dairy wastewater. Water Research 44, 4077–4086. Llorens, M., Pérez, A. B., Aguilar, M. I., Saez, J., Ortuno, J. F. & Meseguer, V. F.  Nitrogen transformation in two subsurface infiltration systems at pilot scale. Ecological Engineering 37, 736–743. Mayo, A. W. & Bigambo, T.  Nitrogen transformation in horizontal subsurface-flow constructed wetlands I: model development. Physics and Chemistry of the Earth 30, 658–667. Metcalf & Eddy  Wastewater Engineering: Treatment and Reuse. 4th edn, McGraw-Hill, New York.

M. M. Pérez et al.

Vertical ﬂow constructed wetlands: kinetics

Molle, P., Lienard, A., Grasmick, A. & Iwema, A.  Effect of reeds and feeding operations on hydraulic behavior of vertical flow constructed wetlands under hydraulic overloads. Water Research 4, 606–612. ONN: NC27/2012  Vertimiento de Aguas Residuales a Las Aguas Terrestres y al Alcantarillado (Wastewater disposal to surface water and sewer), Oficina Nacional de Normalización. La Habana, Cuba. Pérez, M. M.  Evaluación del Comportamiento de Humedales Subsuperficiales de Flujo Vertical en la Purificación de Aguas Residuales Domésticas. Doctor Thesis, Central University of Las Villas, Cuba. Pérez, M. M., Rosa, E., Martínez, P., López, M. E., González, Y. & Monteagudo, M.  Evaluación de la eficiencia de diferentes sustratos de filtros de suelo plantados en la depuración de aguas residuales domésticas (Evaluation of efficiency of planted soil filter substrates in domestic wastewater depuration). CENIC Ciencias Biológicas 40, 99–103. Pérez, M. M., Rosa, E., Tack, F. & Vandemoortel, A.  Caracterización de sustratos de filtros de suelo plantados: Influencia en la depuración de aguas residuales (Characterization of planted soil filters substrate: Influence in wastewater purification). Afinidad 68, 124–128. Pérez, M. M., Rosa, E., Tack, F., Hernández, J. M., Sánchez, R. & Arteaga, L. E.  Vertical subsurface wetlands for wastewater purification. Procedia Engineering 42, 2145–2153. Prochaska, C. A. & Zouboulis, A.  Removal of phosphates by pilot vertical-flow constructed wetlands using a mixture of sand and dolomite as substrate. Ecological Engineering 26, 293–303. Reed, S. C., Crites, R. W. & Middlebrooks, E. J.  Natural Systems for Waste Management and Treatment. McGrawHill, New York, USA. Rosenquit, S. E., Hession, W. C., Eick, M. J. & Vaughan, D. H.  Variability in adsorptive phosphorus removal by structural stormwater best management practices. Ecological Engineering 36, 664–671. Rousseau, D. P. L., Vanrolleghem, P. A. & DePauw, N.  Model based design of horizontal subsurface flow

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constructed treatment wetlands: a review. Water Research 38 (6), 1484–1493. Saeed, T. & Sun, G. Z.  A comparative study on the removal of nutrients and organic matter in wetland reactors employing organic media. Chemical Engineering Journal 171 (2), 439–447. Sikdar, A.  Quantification of Complex Phosphorus Removal Reactions Occurring within Wetlands Filtration Treatment Systems. Doctor of Philosophy Thesis, Department of Civil Engineering, Queen’s University, Kingston, Ontario, Canada. Tang, X. Q., Huang, S. L. & Fciwem, M. S.  Comparison of phosphorus removal between vertical subsurface flow constructed wetlands with different substrates. Water and Environment Journal 23, 180–188. Tietz, A., Kirschner, A., Langergraber, G., Sleytr, K. & Haberl, R.  Characterisation of microbial biocoenosis in vertical subsurface flow constructed wetlands. Science of the Total Environment 380, 163–172. Vymazal, J.  Removal of nutrients in various types of constructed wetlands. Science of the Total Environment 380 (1–3), 48–65. Wang, Y., Zhang, J., Kong, H., Inamori, Y., Xu, K., Inamori, R. & Kondo, T.  A simulation model of nitrogen transformation in reed constructed wetlands. Desalination 235 (1–3), 93–101. Wood, R. B. & McAtamney, C. F.  Constructed wetlands for wastewater treatment: the use of laterite in the bed medium in phosphorus and heavy metal removal. Hydrobiologia 340, 323–331. Yalcuk, J. A. & Ugurlu, A.  Comparison of horizontal and vertical constructed wetland systems for landfill leachate treatment. Bioresource Technology 100, 2521–2526. Yao, F., Shen, G. X., Li, X. L., Li, H. Z., Hu, H. & Ni, W. Z.  A comparative study on the potential of oxygen release by roots of selected wetland plants. Physics and Chemistry of the Earth 36, 475–478. Zhao, Y. J., Hui, Z., Chao, X., Nie, E., Li, H. J. & He, J.  Efficiency of two-stage combinations of subsurface vertical down-flow and up-flow constructed wetland systems for treating variation in influent C/N ratios of domestic wastewater. Ecological Engineering 37 (10), 1546–1554.

First received 23 September 2013; accepted in revised form 2 April 2014. Available online 23 April 2014

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Analysis of extracellular polymeric substance (EPS) release in anaerobic sludge holding tank and its effects on membrane fouling in a membrane bioreactor (MBR) Byeong-Cheol Kim, Duck-Hyun Nam, Ji-Hun Na and Ki-Hoon Kang

ABSTRACT Amongst sludge reduction strategies, the anaerobic side-stream sludge holding tank (SHT) is of particular interest because it has shown significant sludge reduction efficiency. However, due to the anaerobic and starving environment of the SHT, the release of extracellular polymeric substance (EPS) may be stimulated, and it may hamper the application of the SHT to the membrane bioreactor. In order to investigate the effect of sludge storage on EPS release, sludge samples from a pilot-scale sequencing batch reactor coupled with SHT was incubated in a series of bench-scale SHT reactors for different periods of time (0–24 h). The increase in EPS was not significant until 12 h of incubation

Byeong-Cheol Kim Duck-Hyun Nam Ji-Hun Na Ki-Hoon Kang (corresponding author) Technology R&D Institute, Daelim Industrial Co., Ltd, Joonghak-dong, Jongno-gu, Seoul 110–140, Korea E-mail: khkang@daelim.co.kr

(9.3%), while 40.9% of the increase was observed in the sample incubated for 24 h. The rapid increase in EPS concentration after 12 h indicates a greater rate of cell lysis than that with EPS consumption as substrate. Since inducing the initial stage of the endogenous phase within microorganisms is a key factor for the successful operation of the SHT for sludge reduction, the retention time for the SHT should be shorter than the time for the sudden increase in EPS release. Key words

| extracellular polymeric substance, membrane bioreactor, membrane fouling, sludge holding tank

INTRODUCTION The treatment and disposal of waste activated sludge has become more challenging, mainly due to stronger public opposition and stricter regulatory pressure. The London Protocol, which came into force in March 2006 prohibits the ocean dumping of waste sludge. Accordingly, many studies have attempted to minimize excess sludge production by treating sludge using mechanical solubilisation (Nah et al. ), heat treatment (Yan et al. ), chemical oxidation (He & Wei ), or biological hydrolysis (Yang et al. ). However, these methods require the installation of additional unit processes in the sludge handling procedure, and these require a relatively high amount of energy consumption. Among several sludge reduction strategies, the anaerobic side-stream sludge holding tank (SHT) is of particular interest due to its significant sludge reduction efficiency. It also causes no negative effects on sludge settling and effluent water quality (Datta et al. ; Nam & Kang ). The oxic-settling-anaerobic (OSA) process is one of the typical methods utilizing anaerobic SHT installed in the sludge recycling line doi: 10.2166/wst.2014.198

(Chudoba et al. ). Uncoupling metabolism, changes in the ratio of initial substrate to biomass concentration, microbial decay, domination of microorganisms with low growth rate, and involvement of a predator have been proposed as possible mechanisms for yield reduction in the OSA process (Low & Chase ; Chen et al. ). Of these possible solutions, uncoupling metabolism has been widely studied as the most plausible mechanism (Low & Chase ; Chen et al. ). By introducing recycled sludge into anaerobic SHT for a certain period of time, microorganisms are induced into the initial stage of the endogenous growth phase. Once the microorganisms are moved from the SHT to the bio-reactor containing a relatively high amount of substrate (inﬂuent chemical oxygen demand (COD) concentration of about 90–310 mg/L), the inhibition effect of microbial metabolism takes place. This rapid change in the growth environment leads to a sludge yield reduction (Low & Chase ; Chen et al. ). It can also be explained by the reduced production of adenosine triphosphate in microbial cells.

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Analysis of EPS release in anaerobic sludge holding tank and its effects on MBR

Conversely, wastewater reclamation and reuse have been now considered as one of the ways to alleviate water scarcity. As a result, the utilization of membrane in the field of wastewater treatment has become more widespread. For membrane bioreactor (MBR) systems, however, membrane fouling is the most significant obstacle in its application due to declining flux and the need for increased cleaning frequency. Therefore, membrane fouling can increase operating costs. Of the causes of membrane fouling, extracellular polymeric substance (EPS) produced by microorganisms has been recognized as the main source in MBR systems (Chang et al. ; Drews et al. ; Meng et al. ). In particular, for MBR systems, the effects of EPS release in SHT should be considered since the harsh conditions of SHT with little oxygen and substrate may trigger higher amounts of EPS release by microorganisms. Barker & Stuckey () reported that microorganisms release EPS to obtain maintenance energy through endogenous respiration or metabolism of intracellular components when substrate is highly limited. The increased release of EPS, therefore, may act as a barrier to the application of SHT to MBR systems. In this study, 50 m3/day of a pilot sequencing batch reactor (SBR) system, coupled with anaerobic SHT, was operated for about 2 years to quantitatively evaluate the reduction of sludge production. The system’s sludge production was reduced by 63.5% compared to normal SBR. For the evaluation of sludge production, the kinetic parameters estimation method was used as it provides more direct evidence than mixed liquor suspended solids (MLSS) measurement. The detailed results from the experiments have been reported by Nam & Kang (). From the pilot system, settled sludge was taken from the sludge recycling line to SHT, and incubated anaerobically in a series of bench-scale SHT reactors. Using the sludge from each reactor, which had different anaerobic incubation times, EPS was extracted and quantified. In addition, the filterability of each sludge sample was compared using 0.1-μm polyvinylidene fluoride (PVDF) membrane in a stirrer cell. From the experimental results, the appropriateness of SHT application to MBR was evaluated. In addition, optimal design and operation parameters of SHT for MBR systems were suggested.

METHODS Pilot plant operation A pilot-scale SBR process (50 m3/day) coupled with SHT (Figure 1) was operated for about 2 years in a local sewer

Figure 1

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Schematic diagram of SBR pilot plant with SHT .

treatment plant located in the suburban metropolitan area of Seoul, Korea. The SBR system was operated with a 6 h operating cycle, and fed with raw wastewater that was pretreated with 0.2-mm fine screen (Kumsung MC, Korea). For the ‘reaction’ sequence, the reactor was intermittently aerated to operate the SBR as a biological nutrient removal system. On the ‘return’ cycle, 0.5Q (Q ¼ influent flowrate) of settled sludge was pumped into SHT while the sludge incubated in the anaerobic SHT was returned to the SBR reactor by gravitational flow. The SHT was maintained in reductive conditions (oxidation-reduction potential of around 250 mV) with a hydraulic retention time of 6 h, the same as one operation cycle of the SBR system. Detailed descriptions of the design and operation of the pilot SBR system have been reported by Nam & Kang ().

EPS extraction and quantification From the pilot SBR system, settled sludge (with an MLSS concentration of 5,538 ± 18 mg/L, measured using a 0.45-μm filter) was taken from the sludge recycling line to SHT, and incubated anaerobically in a series of bench-scale SHT reactors. Each reactor, with a working volume of 1-L, was sealed and placed in a shaking water bath at 30 C for different periods of time (ranging from 0 to 24 h). After incubation, EPS was extracted and quantified from each sludge sample according to the procedure described by Zhang et al. (). For soluble EPS (S-EPS), 100 mL of each sludge sample was settled for 30 min, and the supernatant was collected for EPS extraction. The settled sludge sample was added, along with deionized water, to the original volume, mixed thoroughly, and centrifuged at 10,000 g for 10 min. Loosely bound EPS (LB-EPS) were extracted from the supernatant. The centrifuged sediment was again added, along with 40 mL of 0.1 N NaOH solution, put in a shaking water bath of 4 C at 300 rpm for 40 min, and centrifuged at W

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Analysis of EPS release in anaerobic sludge holding tank and its effects on MBR

19,000 g for 15 min. The supernatant was neutralized with 0.1 N HCl and used for the extraction of tightly bound EPS (TB-EPS). The total amount of the three EPS fractions was considered as total EPS (T-EPA). Concentrations of carbohydrates and proteins, which are major components of EPS, were analysed by Anthrone (Frølund et al. ) and BCA (Noble & Bailey ) methods, respectively. For the carbohydrate standard, glucose was used while bovine serum albumin was used as the protein standard. All extraction and quantiﬁcation experiments were performed in triplicate.

Filterability test Filterability tests for each sludge sample incubated at different periods of time was carried out at 20 ± 2 C using a stirrer cell of 150 mL (Amicon 8200, Millipore Co., Billerica, MA) and flat sheet PVDF membrane with 0.1-μm pore size and 32 cm2 surface area (VVLP09050, Millipore Co., Billerica, MA), as described in Figure 2. The initial pure water flux, J(0) of the membrane was measured using deionized water. 150 mL of each sludge sample incubated for different periods of time was filtered under constant pressure (0.5 MPa) using N2 gas. After filtration, the cake layer on the membrane surface was carefully removed, and the membrane was backwashed with 150 mL of deionized (DI) water applying the same pressure. Afterwards, membrane flux was again measured using 150 mL of DI water. This cycle was repeated three times. During the filtration, permeate was taken into a reservoir placed on an electronic balance to measure permeate flux. The measured value was automatically recorded every 5 s. In order to investigate the effects of increased EPS concentration on membrane fouling, the reversibility of the flux W

Figure 2

Schematic diagram of ﬁltration test cell.

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by backwashing was calculated using Equations (1) to (3). J(0) Js(nþ1) Js(nþ1) IFn ¼ ¼ 1 J(0) J(0)

RFn ¼

Js(nþ1) Je(n) J(0)

TFn ¼ IFn þ RFn

(1)

(2)

(3)

Js(n) and Je(n) are the starting and ending flux of sludge samples at n-th filtration, respectively. Irreversible fouling, IFn is defined here as accumulated fouling after n-th filtration that is not recovered by backwashing. It is represented as flux at the beginning of the next cycle, Js(nþ1), compared to the initial flux, J(0) (Equation (1)). Reversible fouling, RFn is fouling that is recoverable by backwashing (Equation (2)). Total fouling, TFn is, therefore, the sum of irreversible (IFn) and reversible fouling (RFn) accumulated during n-th filtration (Equation (3)).

RESULTS AND DISCUSSION EPS release in anaerobic SHT In order to investigate the release of EPS in the anaerobic SHT, sludge samples taken from the pilot plant were incubated under reductive anaerobic conditions for 24 h. EPS concentrations were analyzed in three different fractions (soluble, loosely bound, and tightly bound EPS) with respect to incubation time; normalized concentrations with the amount of mixed liquor volatile suspended solids (MLVSS) were also calculated (Table 1). As indicated in Figure 3 and Table 1, the increase in T-EPS was not significant until the incubation time of 12 h (9.3%), while the T-EPS concentration after 24 h steeply increased (about 40.9% relative to the initial concentration). Since EPS can be utilized as substrate by microorganisms, the increase in EPS concentration indicates a greater rate of cell lysis than that for EPS consumption, which, in turn, suggests the initiation of a fully developed endogenous growth phase after 12 h. Although there was no significant change in the ratios of each EPS fraction with incubation time, LB-EPS slightly decreased from 9.7 to 7.4%, while TB-EPS showed a slightly increasing trend from 69.4% to about 74%. According to the ‘unified theory’ proposed by Laspidou & Rittmann (), EPS can be classified into

B.-C. Kim et al.

Table 1

Analysis of EPS release in anaerobic sludge holding tank and its effects on MBR

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Variation of the EPS concentration (in mg/L) with time in the anaerobic SHT

S-EPS

LB-EPS

TB-EPS

HRT

MLSS (mg/L)

MLVSS (mg/L)

C**

T-EPS (mg EPS/g MLVSS)

5520

4140

15.5

12.2

3.7

9.1

82.3

9.6

132.5 (32.0)

5560

4120

13.7

13.0

3.9

9.8

82.3

12.2

134.9 (32.7)

5550

4140

14.3

11.5

3.2

7.1

88.6

16.1

140.8 (34.0)

5520

4140

14.2

9.9

5.5

7.8

90.2

17.1

144.7 (34.9)

5540

4120

29.9

5.8

11.1

2.6

105.2

31.0

185.6 (45.1)

P* ¼ Protein, C** ¼ Carbohydrate.

experiments performed in a similar condition to this experiment (sludge with 5,000 mg/L MLSS at 30 C), the ammonia release rate was increased after 10 h, whereas accelerated EPS release was observed after 12 h. Figure 4 shows the graphical comparison of both experimental results. More than 2 h of the lag time may have been caused by the time required for the EPS release rate surpassing the EPS consumption rate. Although the exact increasing trend of EPS concentrations between 12 and 24 h is not shown in Figure 4 due to the limited numbers of experimental data, it is important to understand that a significant increase in EPS concentration occurs only after initiation of the endogenous phase. Therefore, anaerobic SHT can be incorporated into an MBR system for the purpose of sludge reduction with no significant deterioration of performance on membrane filtration. According to Wang et al. (), however, EPS release is increased at low temperature, while the time to the initial stage of the endogenous phase is delayed. In operation conditions at low temperature, EPS can be released excessively before the initial stage of the endogenous phase, and, W

Figure 3

Variation of EPS concentration in anaerobic SHT with hydraulic retention time (S-EPS: soluble EPS, LB-EPS: loosely bound EPS, TB-EPS: tightly bound EPS, T-EPS: total EPS).

two categories: that is, utilization-associated products formed during substrate utilization and biomass-associated products (BAP) formed from endogenous cell decay. In the absence of substrate supply into the SHT, a decreased ratio of LB-EPS and an increased ratio of TB-EPS may be explained by the increased ratio of BAP caused by the accelerated rate of biomass degradation with time. A signiﬁcant increase in T-EPS after 12 h indicates accelerated cell lysis, which suggests that microorganisms underwent a fully developed endogenous growth phase. In the design phase of the pilot SBR system, the optimal retention time of sludge in SHT was determined by anaerobically incubating settled sludge taken from the same wastewater treatment plant and by monitoring ammonia concentration released from biomass decay with incubation time (Nam & Kang ). Since induction of the initial stage of the endogenous phase within microorganisms is critical for the reduction of sludge production, the time required to accelerate the rate of ammonia release was considered as the optimal retention time of SHT. In the incubation

Figure 4

The increasing rate variation of NH3-N and EPS concentration with time by W

incubation of the settled sludge at 30 C and 5,000 mg/L MLSS conditions.

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Analysis of EPS release in anaerobic sludge holding tank and its effects on MBR

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therefore, the temperature of SHT should be maintained appropriately when it is coupled with an MBR system.

Membrane filtration test Many studies have revealed that EPS produced by microorganisms plays a major role in membrane fouling by analysing foulants on the membrane with Fourier transform infrared (FT-IR) spectroscopy and fluorescence excitationemission matrix (Meng et al. ). Although the amount of EPS released in anaerobic SHT was quantified with respect to retention (or incubation) time, its effect on membrane fouling and reversibility was directly evaluated using a stirrer cell with 0.1-μm PVDF flat sheet membrane. Some flux data measured under the constant pressure (0.5 MPa) are shown in Figure 5. As shown in the figure, filtration curves of the samples incubated up to 12 h exhibited no significant differences. By backwashing with deionized water after each filtration of the sample, the flux recovered was greater than 90 and 80%, respectively, both with the samples incubated for 0 and 12 h. Conversely, the filtration curve of the sample incubated for 24 h indicated a more rapid decrease in flux and less recovery by backwashing. After the first backwashing, the flux recovered was about 78%, and after the second backwashing the recovery was limited to about 59%. Using the filtration data, reversibility of the fouling was calculated using Equations (1) to (3), and the results are shown in Figure 6. In tests with the samples incubated up to 12 h, irreversible fouling was in the range of 16–18% in the first filtration, and no considerable increase in irreversibility was observed in the repeated filtration with backwashing (up to 21%). However, in tests with the

Figure 6

Reversibility of the membranes by backwashing after ﬁltrations.

sample incubated for 24 h, irreversible fouling in the first filtration was about 27%, and increased to 35% at the third filtration. The decreased reversibility of the sample incubated for longer than 12 h indicates the importance of proper design of SHT in terms of hydraulic retention time. If the retention time is longer than the time to the initial stage of the endogenous phase, the EPS release rate will surpass the uptake rate. The accumulated EPS will increase irreversible membrane fouling which cannot be recovered by physical backwashing. Therefore, operation of SHT with a longer retention time may create the need for more frequent chemical cleaning, higher operating costs, and ultimately, a shorter membrane lifetime. The increased irreversible fouling was also visually identified by taking scanning electron microscopy (SEM) images as shown in Figure 7. After the filtration tests, membranes were backwashed again before taking images. In the membrane used for the filtration of the 24-h sample, more significant fouling was observed compared to other membranes.

CONCLUSION

Figure 5

Filtration curves of the sludge samples incubated for 0 to 24 h in SHT using 0.1-μm MF membrane at constant pressure (0.5 MPa).

In this study, the release of EPS and its effect on membrane filterability was evaluated with respect to anaerobic incubation time in SHT reactors. Within 12 h of anaerobic incubation, no significant increase in EPS concentration was observed (9.3%), while a 40.9% increase was observed after 24 h. In the membrane filtration test, irreversible fouling was not significantly increased in the sludge samples incubated for 12 h (up to 21% with repeated filtration), but it increased to about 35% for the sample incubated for 24 h. According to the findings of this study,

Figure 7

B.-C. Kim et al.

Analysis of EPS release in anaerobic sludge holding tank and its effects on MBR

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SEM images of the membranes after three times ﬁltration of the MLSS.

understanding EPS release characteristics and time is very important for optimal design and operation of SHT. In particular, when applied to MBR systems, preliminary experiments should be performed at a given operation temperature.

REFERENCES Barker, D. J. & Stuckey, D. C.  A review of soluble microbial products (SMP) in wastewater treatment systems. Water Res. 33 (14), 3063–3082. Chang, I. S., Bag, S. Q. & Lee, C. H.  Effects of membrane fouling on solute rejection during membrane ﬁltration of activated sludge. Process Biochem. 36 (8), 855–860. Chen, G. H., An, K. J., Saby, S., Brois, E. & Djafer, M.  Possible cause of excess sludge reduction in an oxic-settlinganaerobic activated sludge process (OSA process). Water Res. 37 (16), 3855–3866. Chudoba, P., Chudoba, J. & Capdeville, B.  The aspect of energetic uncoupling of microbial growth in the activated sludge process: OSA system. Water Sci. Technol. 26 (9–11), 2477–2480. Datta, T., Liu, Y. & Goel, R.  Evaluation of simultaneous nutrient removal and sludge reduction using laboratory scale sequencing batch reactors. Chemosphere 76 (5), 697–705. Drews, A., Lee, C. H. & Kraume, M.  Membrane fouling – a review on the role of EPS. Desalination 200 (1–3), 186–188.

Frølund, B., Palmgren, R., Keiding, K. & Nielsen, P. H.  Extraction of extracellular polymers from activated sludge using a cation exchange resin. Water Res. 30 (8), 1749–1758. He, M. H. & Wei, C. H.  Performance of membrane bioreactor (MBR) system with sludge Fenton oxidation process for minimization of excess sludge production. J. Hazard Mater. 176 (1–3), 576–601. Laspidou, C. S. & Rittmann, B. E.  A unified theory for extracellular polymeric substances, soluble microbial products, and active and inert biomass. Water Res. 36 (11), 2711–2720. Low, E. W. & Chase, H. A.  Review paper: reducing reduction of excess biomass during wastewater treatment. Water Res. 33 (5), 1119–1132. Meng, F., Chase, S., Drews, A., Kraume, M., Shin, H. & Yang, F.  Recent advances in membrane bioreactors (MBRs): Membrane fouling and membrane material. Water Res. 43 (6), 1489–1512. Nah, L. W., Kang, Y. W., Hwang, K. Y. & Song, W. K.  Mechanical pre-treatment of waste activated sludge for anaerobic digestion process. Water Res. 34 (8), 2362–2368. Nam, D.-H. & Kang, K.-H.  Evaluation of WAS reduction efficiency using kinetic parameters in pilot-scale SBR operated with anaerobic sludge holding tank. Water Sci. Technol. 67 (12), 2838–2844. Noble, J. E. & Bailey, M. J. A.  Quantitation of protein. Method. Enzymol. 463, 73–95. Wang, Z., Song, Y., Mei, L., Jinmiao, Y., Yuanyuan, W., Zhang, H. & Zhaohui, L.  Experimental study of filterability

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Analysis of EPS release in anaerobic sludge holding tank and its effects on MBR

behavior of model extracellular polymeric substance solutions in dead-end membrane ﬁltration. Desalination 249 (3), 1380–1384. Yan, S., Miyanaga, K., Xing, X.-H. & Tanji, Y.  Succession of bacterial community and enzymatic activities of activated sludge by heat-treatment for reduction of excess sludge. Biochem. Eng. J. 39 (3), 598–603.

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Yang, X. F., Xie, M. L. & Liu, Y.  Metabolic uncouplers reduce excess sludge production in an activated sludge process. Process Biochem. 38 (9), 1373–1377. Zhang, B., Sun, B., Jin, M., Gong, T. & Gao, Z.  Extraction and analysis of extracellular polymeric substances in membrane fouling in submerged MBR. Desalination 227 (1–3), 286–294.

First received 4 November 2013; accepted in revised form 9 April 2014. Available online 22 April 2014

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Agar-encapsulated adsorbent based on leaf of platanus sp. to adsorb cadmium ion in water Eko Siswoyo, Nozomi Endo, Yoshihiro Mihara and Shunitz Tanaka

ABSTRACT A low cost and environmentally friendly adsorbent was developed based on leaf of platanus sp. to adsorb cadmium ion in water. The adsorbent was modified with citric acid and then also encapsulated in agar for easy separation after the adsorption process. Parameters such as adsorbent dose, stirring time, solution pH and modification of the adsorbent with citric acid were investigated in a batch experiment in order to determine the optimum condition for Cd (II) adsorption. Based on the Langmuir isotherm adsorption model, the adsorption capacity of cadmium ion for raw adsorbent, modified adsorbent with citric acid and encapsulated adsorbent were 3.69, 15.31 and 6.89 mg/g, respectively. The high adsorption capacity after treatment with citric acid may be due to the increase in carboxylic content of the adsorbent surface and also the increase of surface area and pore volume of the adsorbent. With this high adsorption capacity for cadmium ion and an abundance of raw materials, this bio-adsorbent could be considered as a low cost adsorbent in the near future. Key words

Eko Siswoyo (corresponding author) Nozomi Endo Yoshihiro Mihara Shunitz Tanaka Graduate School of Environmental Science, Division of Environmental Science Development, Hokkaido University, Sapporo Kitaku North 10 West 5, Hokkaido 060-0810, Japan E-mail: ekosiswoyo@ees.hokudai.ac.jp Eko Siswoyo Department of Environmental Engineering, Faculty of Civil Engineering and Planning, Islamic University of Indonesia, Jl. Kaliurang Km 14.4, Yogyakarta, Indonesia

| cadmium ion, citric acid, encapsulated adsorbent, low cost adsorbent, platanus leaf

INTRODUCTION The issue of environmental contaminations with heavy metals is considered a serious environmental problem because of its potential for damage to human health and the environment. In the environment, the presence of cadmium, one of the most toxic metal ions, sometimes leads to a serious problem for human beings and the ecosystem. Cadmium can be released to the environment through many industrial activities such as ceramics, metal plating, textiles, etc. (Rao et al. ; Wang et al. ). Itai-itai disease was caused by cadmium poisoning, which resulted in softening of the bone and kidney failure in the Jinzu River area of the Toyama Prefecture, was one of the most severe environmental events in Japan. The rate of death among patients was 72.6% and it is considered that itai-itai disease has a long-lasting negative effect on a patient’s life span (Kawano et al. ). Many technologies have been employed in order to minimize the pollution problem of heavy metals; for example, membrane technology, ion exchange, phytoremediation, adsorption, etc. Adsorption is one of the common methods that have been widely used for water and wastewater treatment. As an adsorbent, activated carbon has often been utilized in many countries; however, this material is quite expensive. The high cost of activated doi: 10.2166/wst.2014.190

carbon has inspired many studies into the development of alternative low cost adsorbent materials. Recently, some researchers have been attracted to the development of bio-adsorbents based on the leaves of some plants due to the high removal capacity for adsorbing heavy metal ions from solution. For example, Pinus roxburghii leaves were utilized to remove heavy metals from industrial effluent (Tewari & Vivekanand ). In order to increase the adsorption capacity, bio-adsorbents were modified with some chemicals. Li et al. () reported that modification of Imperata cylindrical leaf powder using sulphuric acid and phosphoric acid resulted in high adsorption capacity for nickel in solution. Chemical modification on Moringa oleifera leaf powder using NaOH followed by citric acid was applied for the adsorption of Cd (II), Cu (II) and Ni (II) (Reddy et al. ). In Japan and many other northern hemisphere countries, the platanus tree is abundant and often used as ornamental and roadsides trees. The fallen leaves of this tree are usually collected and disposed of as garbage in Japan. The leaf of this tree has potential as a low cost and environmentally friendly adsorbent material; however, the study of the utilization of the leaf of the platanus tree as an adsorbent material is still limited.

E. Siswoyo et al.

Encapsulated adsorbent in agar based on leaf of platanus sp. to adsorb cadmium ion in water

The purpose of this study was to investigate the adsorption capacity of an adsorbent encapsulated in agar based on the leaf of platanus sp. for the removal of cadmium ion in water. Powdered platanus leaf was treated with citric acid and then encapsulated in agar as a binder material for easy separation of the adsorbent from solution after the process of adsorption of cadmium ion.

MATERIALS AND METHODS Preparation of adsorbent and stock solution Adsorbents were prepared based on leaf of platanus sp. collected from the Hokkaido University campus area. The dry leaf was rinsed with tap water, then washed continuously with distilled water to remove dirt and other particulate matter, dried at 80 C for 24 hours and then crushed into powder form. In order to increase the adsorption capacity of the adsorbent, 1.3 M of citric acid was utilized as an activator agent with a composition of 1 g (adsorbent): 4 mL (citric acid) and then heated at 150 C for 30 min. The activated adsorbent then was encapsulated in agar with the ratio of 1 g (agar): 2.5 g (activated adsorbent). The stock solution of cadmium ion (Cd2þ) was prepared with Cd(NO3)2.4H2O from WAKO Pure Chemical Co. (Osaka, Japan). The standard solutions for analysis of Cd2þ by atomic adsorption spectrophotometer (AAS HITACHI A-2000, Hitachi, Japan) were prepared by using a pure standards solution from WAKO Pure Chemical Co. (Osaka, Japan). W

Characterization of adsorbent In order to obtain the surface area of the adsorbent, scanning electron microscopy/energy dispersive spectroscopy (SEM/ EDS) (JEOL JSM-6360 LA, Japan) was employed. Furthermore, the elemental test was undertaken using an elemental instrument (MICRO CORDER JM10, Yanaco, Japan) in order to determine the amounts of C, H, N, O, S and ash of the adsorbent material. Fourier transform infrared spectroscopy (FTIR) (FT/IR-4100) and Brunauer-Emmett-Teller (BET) (BELSORP-mini, BEL Japan Inc., Osaka, Japan) were utilized in order to determine the functional groups, surface area and pore volume of adsorbent materials. Adsorption experiment The adsorption process in the present study was conducted using a batch system. Parameters such as mass of adsorbent,

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pH of the solution, stirring time and initial concentration of cadmium ion in the solution were investigated in order to determine the ability and the optimum condition for adsorption using this adsorbent. In order to estimate the effect of the mass of the adsorbent, 25–200 mg of adsorbent was added to 50 ml of cadmium solution and agitated at 1,000 rpm for 2 hours using a shaker (EYELA Cute Mixer CM-1000). The influence of pH on the adsorption of cadmium ion was investigated by using a solution of pH 2 to 8. Acetic buffers, HNO3 and NaOH were utilized to adjust the desired pH of the solution. Various stirring times from 15 to 1,440 min were applied in order to determine the influence of stirring time on the adsorption of cadmium ion. After reaching equilibrium, the solution was centrifuged at 4,000 rpm for 5 min using a centrifuge (IEC61010-2-020, KUBOTA, Japan) and then the concentration of cadmium ion in the supernatant solution was determined using a flame atomic absorption spectrophotometer (HITACHI A-2000, Japan).

RESULTS AND DISCUSSION Properties of adsorbent Table 1 shows the composition of C, H, O, N, S and ash of adsorbent materials before and after being activated with citric acid. The amount of O in the adsorbent material signiﬁcantly increased from 23.03 to 40.22% and ash decreased from 20.11 to 1.79% after treatment with citric acid. The SEM photograph of adsorbent materials before and after being activated with citric acid are shown in Figure 1. It can be seen that the surface of the adsorbent after modiﬁcation with citric acid appears to be more porous compared to raw adsorbent. Treatment of the adsorbent material with citric acid resulted in more pores on the surface of the adsorbent. Table 1

Elemental analysis of adsorbent materials before and after treatment with citric acid

Adsorbent Element (%w/w)

Raw adsorbent

After treatment with citric acid

Carbon (C)

49.90

Hydrogen (H)

5.78

6.14

Oxygen (O)

23.03

40.22

Nitrogen (N)

1.10

1.95

Sulphate (S)

Ash

20.11

1.79

Figure 1

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Encapsulated adsorbent in agar based on leaf of platanus sp. to adsorb cadmium ion in water

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SEM photograph of adsorbent. (a) Raw adsorbent. (b) After being activated with citric acid.

Based on the data obtained by BET analysis, the surface area and volume of adsorbent pores before and after modiﬁcation with citric acid are 19.28 m2 g 1 and 4.43 cm3 g 1, and 31.81 m2 g 1 and 7.31 cm3 g 1, respectively. These data support the SEM photograph in that modiﬁcation with citric acid resulted in more pores on the surface of the adsorbent. FTIR analysis was employed in order to identify the characteristic functional groups on the surface of adsorbents. Figure 2 shows the FTIR spectra of the adsorbents before and after being activated with citric acid. The FTIR spectrum of raw adsorbent and activated adsorbent showed a broad peak at 3,400 cm 1, indicating the presence of macromolecular compounds such as cellulose, lignin, pectin, etc. The double peaks that appeared in both spectra at wave numbers 2,922 and 2,849 cm 1 may be due to the asymmetric and symmetric stretch of aliphatic chains (-CH), respectively. The peak around

Figure 2

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FTIR spectra of adsorbent before and after treatment with citric acid.

1,435–1,251 cm 1 indicated the characteristic main skeleton of cellulose (Reddy et al. ). The peak around 3,371 cm 1 indicated the presence of carboxylic O-H groups, and the peaks around 1,067 cm 1 indicated C ¼ O group of primary hydroxyl stretching, which may be attributed to the cellulose structure of platanus (Nguyen & Ha ). The activated adsorbent has a strong characteristic stretching vibration absorption band of carboxyl group at 1,719 cm 1 as an effect of treatment with citric acid. Based on the data obtained by SEM, FTIR and BET analysis, it can be concluded that modiﬁcation with citric acid resulted in physical and chemical properties on the surface of the adsorbent and increased the adsorption capacity of cadmium in water. The increase of the sorption of metal ions may be due to the esteriﬁcation process, which also increases the carboxylic content of the surface of the adsorbents (Low et al. ).

E. Siswoyo et al.

Encapsulated adsorbent in agar based on leaf of platanus sp. to adsorb cadmium ion in water

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Effect of dose of adsorbent on the adsorption of cadmium ion The effect adsorbent dose is shown in Figure 3. For 10 mg/L of initial cadmium concentration, the adsorption capacity for all mass of adsorbent from 25 to 200 mg was almost constant; therefore, 25 mg was favorable as an ideal mass for adsorbent. When the initial concentration of cadmium increased, the removal efﬁciency of adsorption increased linearly by the increase of adsorbent mass and became constant from 150 mg. This result proved that the adsorption of cadmium ion in water was absolutely due to the presence of the adsorbent. Effect of pH

Figure 4

The inﬂuence of pH on the adsorption of cadmium (mass of adsorbent: 100 mg, volume of Cd soln.: 50 ml, Cd initial.: 10 mg/L, and stirring time: 2

2þ

The binding of cadmium ion (Cd ) with the surface functional groups of adsorbent strongly depended on the solution pH. Figure 4 shows the effect of pH on the adsorption of cadmium ion in water. The removal efficiency of cadmium ion increased with the increase of pH because cadmium ion forms a complex with some acidic functional groups in the adsorbent (Khalfa & Bagane ). The removal of metal cation at any pH was much greater than that by hydroxide precipitation. Adsorption of metal cation on adsorbent was influenced by the nature of the adsorbent surface and the distribution of metal species, which also depends on the pH of the solution. The removal efficiency of cadmium decreased with the decrease of pH because protons compete with metal ion for the binding sites on the adsorbent surface. Furthermore, the protons decrease negative charges by association of the functional group with protons (Siswoyo & Tanaka ). The increase in the removal of metal ions as pH increases can be explained on the basis of the decrease in Hþ on the surface, which results in less repulsion than with metal ions (Namasivayam & Ranganathan ; Kadirvelu & Namasivayam ;

hours).

Souag et al. ). In the highly acidic medium, the dissolution of the adsorbent appears to consequently decrease in the active sites of the adsorbent. The highly protonated adsorbent surface is not favorable for the uptake of cadmium ion because of electrostatic repulsion (Singh et al. ; Papandreou et al. ). It is known that the precipitation of cadmium ion in solution starts at around pH 8.5 (Lee & Davis ; Wang et al. ; Khosravan & Lashkari ). The recent study conﬁrmed that cadmium ions were precipitated at pH 9, and pH 6 to 8 was ideal for adsorption of cadmium ion using this adsorbent. Effect of stirring time The effect of stirring time on the adsorption of cadmium ion is shown in Figure 5. The equilibrium adsorption for powder and encapsulated adsorbent on 10 mg/l of cadmium was achieved after around 30 and 1,440 min, respectively. The equilibrium time of the encapsulated adsorbent was longer than the powder adsorbent because the binding site of the encapsulated adsorbent is covered by agar, and the quick equilibrium time is due to the particle size (Messaouda et al. ). This result is important because the equilibrium time is one of the considerations for economical application in wastewater treatment plants (Kadirvelu & Namasivayam ; Rao et al. ). Adsorption isotherm

Figure 3

Effect of dose of adsorbent on cadmium adsorption (volume of Cd soln.: 50 ml, Cd initial.: 10 mg/L, pH of soln.: 6, and stirring time: 2 hours).

The Langmuir and Freundlich isotherm models were applied to the equilibrium constant of adsorption by the following equations.

E. Siswoyo et al.

Figure 5

Encapsulated adsorbent in agar based on leaf of platanus sp. to adsorb cadmium ion in water

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The inﬂuence of stirring time on the adsorption of cadmium (mass of adsorbent: 100 mg volume of Cd soln.: 50 ml, pH of soln.: 6.0, and Cd initial:

Figure 6

10 mg/L).

Langmuir isotherm for adsorption of Cd(II) (mass of adsorbent: 100 mg, volume of Cd soln.: 50 ml, pH of soln.: 6.0, and stirring time: 2 hours and 24 hours for encapsulated adsorbent).

The equation of the Langmuir isotherm model Ce 1 1 ¼ Ce þ KL :qm qe qm

(1)

where Ce is the equilibrium concentration (mg/l), qe is the adsorbed amount of cadmium at equilibrium (mg/g), qm and KL are the Langmuir constants related to the adsorption capacity and energy of adsorption, respectively. From the equation above, a plot of Ce/qe versus Ce will be used to determine the values of qm and KL as the tangent and intercept of the curve. The equation of the Freundlich isotherm model qe ¼ Kf Ce1=n

(2)

where qe is the amount of adsorbed (mg/g), Ce is the equilibrium concentration (mg/l) and Kf and n are constants. To determine the amount of Kf and n, plots between log Ce and log qe

Logðqe Þ ¼ log Kf þ 1=n logðCe Þ

RL ¼ 1=ð1 þ KL C0 Þ

(4)

where KL is the Langmuir constant and C0 is the initial concentration of cadmium ion in the solution. The RL value indicates the shape of the isotherm as follows:

RL > 1

Unfavorable

RL ¼ 1

Linear

0 < RL < 1

Favorable

RL ¼ 0

Irreversible

(3)

The Langmuir adsorption isotherm is shown in Figure 6 and the result of the Langmuir and Freundlich isotherm Table 2

models of inactivated, activated and encapsulated adsorbents determined from the above equations are shown in Table 2. The essential characteristics of the Langmuir isotherm can be expressed due to a dimensionless constant of the separation factor or equilibrium parameter, RL, which is deﬁned as

The result of Langmuir and Freundlich isotherm models

Langmuir model

Freundlich model

The RL value between 0 and 1 is favorable adsorption (McKay et al. ). The RL values of raw, modiﬁed and encapsulated adsorbents for the concentrations of 10 mg/L Cd2þ were 0.016, 0.069 and 0.039, respectively. It means that the bio-adsorbents based on platanus leaf were favorable for the adsorption of cadmium ion in water.

qm KL(l/mg)

3.69

6.29

0.978 4.403 2.61 0.972

Activated with citric acid

15.31

1.36

0.939 6.725 3.27 0.993

Encapsulated in agar

6.89

2.48

0.948 4.709 2.64 0.991

Adsorbent

Inactivated

(mg/g)

CONCLUSION The present study shows that bio-adsorbents based on platanus leaf modiﬁed with citric acid achieve high efﬁciency on

E. Siswoyo et al.

Encapsulated adsorbent in agar based on leaf of platanus sp. to adsorb cadmium ion in water

the adsorption of cadmium ion in water. The batch adsorption experiments have been conducted in different initial cadmium ion concentration, dose of adsorbent, contact time and solution pH. The adsorption of cadmium ion was strongly pH-dependent and pH 6 to 8 was favorable for the adsorbent. Encapsulation of the bio-adsorbent in agar decreased the adsorption capacity of cadmium ion; however, the adsorbent material can be separated easily after the adsorption process. Finally, with this high efﬁciency, easyto-separate, low cost and environmentally friendly process for the removal of cadmium ion in water, the adsorbents could be considered as a promising solution for water and waste water treatment in the near future.

REFERENCES Kadirvelu, K. & Namasivayam, C.  Activated carbon from coconut coirpith as metal adsorbent: adsorption of Cd(II) from aqueous solution. Advances in Environmental Research 7, 471–478. Kawano, S., Nakagawa, H., Okumura, Y. & Tsujikawa, K.  A mortality study of patients with Itai-itai disease. Environmental Research 40, 98–102. Khalfa, L. & Bagane, M.  Cadmium removal from aqueous solution by a Tunisian smectitic natural and activated clay: thermodynamic study. Journal of Encapsulation Absorption Science 1, 65–71. Khosravan, A. & Lashkari, B.  Adsorption of Cd(II) by dried activated sludge. Iranian Journal of Chemical Engineering 8, 41–56. Lee, S. M. & Davis, A. P.  Removal of Cu(II) and Cd(II) from aqueous solution by seafood processing waste sludge. Journal of Water Research 35, 534–540. Li, Z., Teng, T. T., Alkarkhi, A. F. M., Rafatullah, M. & Low, L. W.  Chemical modiﬁcation of Imperata cylindrica leaf powder for heavy metal ion adsorption. Water Air Soil Pollution 224:1505, 1–14. Low, K. S., Lee, C. K. & Mak, S. M.  Sorption of copper and lead by citric acid modiﬁed wood. Wood Science and Technology 38, 629–640.

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McKay, G., Blair, H. S. & Gardener, J. R.  Adsorption of dyes on chitin. I. Equilibrium studies. Journal of Application Polymer Science 27, 3043–3057. Messaouda, S., Larouci, M., Meddah, B. & Velemens, P.  The sorption of lead, cadmium, copper and zinc ions from aqueous solutions on a raw diatomite from Algeria. Water Science and Technology 65 (10), 1729–1737. Namasivayam, C. & Ranganathan, K.  Removal of Cd(II) from wastewater by adsorption on waste Fe(III)/Cr(III) hydroxide. Water Research 29, 1737–1744. Nguyen, D. T. & Ha, L. N.  Cellulose modified with citric acid and its absorption of Pb2þ and Cd2þ ions. Proceeding of International Electronic Conference on Synthetic Organic Chemistry (ECSOC-13). Papandreou, A., Stournaras, C. J. & Panias, D.  Copper and cadmium adsorption on pellets made from fired coal fly ash. Journal of Hazardous Materials 148, 538–547. Rao, M. M., Ramesh, A., Rao, G. P. C. & Seshaiah, K.  Removal of copper and cadmium from the aqueous solutions by activated carbon derived from Ceiba petandra hulls. Journal of Hazardous Materials B 129, 123–129. Reddy, D. H. K., Seshaiah, K., Reddy, A. V. R. & Lee, S. M.  Optimization of Cd(II), Cu(II) and Ni(II) biosorption by chemically modified Moringa oleifera leaves powder. Carbohydrate Polymers 88, 1077–1086. Singh, D. B., Rupainwar, D. C., Prasad, G. & Jayaprakash, K. C.  Studies on the Cd(II) removal from water by adsorption. Journal of Hazardous Materials 60, 29–40. Siswoyo, E. & Tanaka, S.  Development of eco-adsorbent based on solid waste of paper industry to adsorb cadmium ion in water. Journal of Clean Energy Technology 1 (3), 198–201. Souag, R., Touaibia, D., Benayada, B. & Boucenna, A.  Adsorption of heavy metals (Cd, Zn and Pb) from water using keratin powder prepared from Algerian sheep hoofs. European Journal of Scientific Research 35, 416–425. Tewari, H. & Vivekanand.  Removal of heavy metals from industrial effluent using Pinus roxburghii leaves as biosorbent: equilibrium modeling. Water Science and Technology 67 (9), 1894–1900. Wang, Y., Tang, X., Chen, Y., Zhan, L., Li, Z. & Tang, Q.  Adsorption behavior and mechanism of Cd(II) on loess soil from China. Journal of Hazardous Materials 172, 30–37.

First received 26 November 2013; accepted in revised form 7 April 2014. Available online 23 April 2014

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Analysis of nitrogenous and algal oxygen demand in efﬂuent from a system of aerated lagoons followed by polishing pond Hassan Khorsandi, Rahimeh Alizadeh, Horiyeh Tosinejad and Hadi Porghaffar

ABSTRACT In this descriptive-analytical study, nitrogenous and algal oxygen demand were assessed for effluent from a system of facultative partially mixed lagoons followed by the polishing pond using 120 grab samples over 1 year. Filtered and non-filtered samples of polishing pond effluent were tested in the presence and absence of a nitrification inhibitor. Effective factors, including 5-day biochemical and chemical oxygen demand (BOD and COD), total suspended solids (TSS), dissolved oxygen, chlorophyll A, and temperature, were measured using standard methods for water and wastewater tests. The results were analyzed using repeated measures analysis of variance with SPSS version 16. Findings show that the annual mean of the total 5-day BOD in the effluent from the polishing pond consisted of 44.92% as the algal carbonaceous biochemical oxygen demand (CBOD), 43.61% as the nitrogenous biochemical oxygen demand (NBOD), and 11.47% as the soluble CBOD. According to

Hassan Khorsandi (corresponding author) Assist. Prof. of Social Determinants of Health Research Center and Environmental Health Department, School of Public Health, Urmia University of Medical Sciences, Urmia, Iran E-mail: hassankhorsandi@yahoo.com Rahimeh Alizadeh Horiyeh Tosinejad Hadi Porghaffar Municipal Water and Wastewater Company of West Azerbaijan province, Urmia, Iran

this study, the annual mean ratios of algal COD and 5-day algal CBOD to TSS were 0.8 and 0.37, respectively. As the results demonstrate, undertaking quality evaluation of the final effluent from the lagoons without considering nitrogenous and algal oxygen demand would undermine effluent quality assessment and interpretation of the performance of the wastewater treatment plant. Key words

| aerated lagoon, algae, biochemical oxygen demand, nitriﬁcation, polishing pond, wastewater

INTRODUCTION The total biochemical oxygen demand (TBOD) of wastewater is the sum of carbonaceous biochemical oxygen demand (CBOD) and nitrogenous biochemical oxygen demand (NBOD) (Gerardi ; Vesilind et al. ; Bitton ). Unlike carbonaceous oxygen demand, which is proportional to the concentration of biodegradable carbonaceous organic compounds, the nitrogenous oxygen demand exerted during the 5-day test is proportional to the number of initial nitrifying organisms present in the sample, as well as to the availability of ammonia (Rich ). Typically, due to the slow growth of nitriﬁers, it is assumed that the nitrogenous oxygen demand (NOD) does not affect the 5-day biochemical oxygen demand (BOD5) experiment (Tchobanoglous et al. ; Bitton ). Although this assumption is valid on untreated wastewater, it is not correct for treated wastewater, especially for efﬂuent doi: 10.2166/wst.2014.194

from the system of aerated lagoons followed by polishing pond, because the 5-day BOD in effluent from aerated lagoons often contains NBOD (Mara & Pearson ; Mara ). Therefore, the US Environmental Protection Agency authorized the use of a nitrification inhibitor in the BOD5 test and using the CBOD5 term, instead of using the BOD5, for quality assessment of a secondary treatment effluent especially in aerated lagoons and stabilization ponds (Barth ; Dague ; Mara & Pearson ). According to EPA recommendations, the average 30-day CBOD5 for secondary treatment effluent should not exceed 25 mg/L (Tchobanoglous et al. ). However, due to high variations in lagoon and pond effluent NBOD (5–50 mg/L), measuring BOD5 irrespective of NBOD influence would cause misinterpretation of the results (Mara & Pearson ; Mara ).

H. Khorsandi et al.

Nitrogenous and algal oxygen demand in efﬂuent from lagoon

Hall & Foxen () found that nitrification was the cause of 60% of violations in the maximum allowable BOD5 in the aerated lagoon effluent. The interference of nitrification on BOD5 has been studied and confirmed by various researchers; for example, Dague (), Barth () and Chapman et al. (). Conversely, a significant portion of CBOD5 measured in the lagoon and pond effluent is due to the growth and propagation of algae, while the influent streams do not contain any algae (Mara & Pearson ; Ramaraj et al. ). Algae are photosynthetic organisms containing chlorophyll that can be found in water, air, soil and on plants. Algae growth is directly affected by the concentration of nitrogen and phosphorus in water resources (Griffiths ; Ramaraj et al. ). Wastewater contains excessive amounts of nitrogen and phosphorus (Tchobanoglous et al. ; Khorsandi et al. a, b), therefore, providing suitable conditions for algae growth and proliferation due to the long retention time in the lagoons (Yang et al. ). A substantial proportion of BOD5 and chemical oxygen demand (COD) in aerated lagoon effluent is due to interference from the presence of algae, since algae make up more than 80% of total suspended solids (TSS) in effluent from aerated lagoons (Mara ; Gronlund et al. ). Rich () investigated the interference caused by algae on TSS, COD and BOD5 parameters in effluent from aerated lagoon systems and concluded that the standard limit of 25 mg/L for CBOD5 and 30 mg/L for TSS is rarely reached due to the growth of algae. The effects of algal and nitrogenous oxygen demand on the effluent quality of polishing ponds have not been studied fully and simultaneously; furthermore, their roles are often overlooked in the interpretation of effluent quality in wastewater treatment plants (WWTPs). Therefore, the aim of this study is, for the first time in Iran, to investigate the analysis of nitrogenous and algal oxygen demand in effluent from the system of aerated lagoons followed by polishing pond.

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algae growth and nitrification occurrence in the lagoons and polishing pond. The characteristics and the flowsheet for the studied lagoons system are illustrated in Table 1 and Figure 1, respectively. During the study, 120 grab samples were collected from the polishing pond effluent with time intervals of 7 to 10 days at 08.00, 12.00 and 18.00. All tests were performed according to Standard Methods for the Examination of Water and Wastewater (APHA ). Dissolved oxygen (DO) and pH measurements were carried out using a YSI 55 DO meter (YSI Company Inc., USA) and a Schott pH meter model CG-824 (Schott UK Ltd, UK), respectively. COD experiments were analyzed via the closed reflux-colorimetric method using a Spectonic 20 D spectrophotometer (Thermo Fisher Scientific Inc., USA). BOD measurements were determined using an OxiTop respirometer (WTW GmbH, Germany). After extracting chlorophyll A with acetone, it was measured using the 10200 H spectrophotometric method. A few drops of effluent including algae were microscopically observed by Leica DM500 microscope (HACH, Germany). Allylthiourea (HACH, Germany) was used as a nitrification inhibitor, and the samples were filtered through M&N 89/90 filters (Macherey-Nagel GmbH & Co. KG, Germany) with 0.5-micrometer pore size. To determine algal COD, COD tests were conducted on filtered and non-filtered samples. For studying the interference of nitrogenous and algal BOD on the quality assessment of effluent, BOD5 tests were performed under the following conditions, and results were compared by repeated measures analysis of variance (RMANOVA) using SPSS version 16 software: 1. Filtered sample, in the absence of nitrification inhibitor. 2. Filtered sample, in the presence of nitrification inhibitor. 3. Non-filtered sample, in the absence of nitrification inhibitor. 4. Non-filtered sample, in the presence of nitrification inhibitor. Table 1

Characteristics of facultative partially mixed lagoons and polishing pond in the Miandoab WWTP, Iran

METHODS OLR

This descriptive-analytical study was implemented in the Miandoab WWTP, Iran, from July 2009 to June 2010. Miandoab WWTP consists of three facultative partially mixed lagoons in series followed by a maturation pond, which acts as a polishing pond. Due to the incompleteness of the wastewater collection network, the process is working with 57% of its ﬁnal capacity, causing an increase in total hydraulic retention time from 14.57 to 25.3 days and in

Lagoon/pond

Length m

Width m

Depth m

HRT day

Kg TBOD/ha.d

F.L1-1

150

100

3.2

7.7

1,280

F.L1-2

150

100

3.2

7.7

1,280

F.L2

300

100

3.2

7.8

F.L3

300

100

3.2

7.8

Polishing pond

165

100

1.5

560

F.L ¼ facultative partially mixed lagoons, TBOD ¼ NBOD þ CBOD.

Figure 1

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Nitrogenous and algal oxygen demand in efﬂuent from lagoon

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Flowsheet of facultative partially mixed lagoons followed by polishing pond in the Miandoab WWTP, Iran.

Comparison of the BOD5 for non-filtered samples in the absence of nitrification inhibitor (nfBOD5-nATU) and in the presence of nitrification inhibitor (nfBOD5-ATU) illustrates the impact of partial nitrification on BOD5 in the presence of algae. Comparison of the BOD5 for filtered samples in the absence of nitrification inhibitor (fBOD5-nATU) and in the presence of nitrification inhibitor (fBOD5-ATU) shows the effect of partial nitrification on BOD5 in the absence of algae (Stander & Theodore ). Comparison of the BOD5 for non-filtered samples in the presence of inhibitor (nfBOD5-ATU) with the BOD5 for filtered samples containing inhibitor (fBOD5-ATU) indicates the impacts of algae on BOD5 in the absence of nitrifiers. Because of the low prevalence of non-algal suspended solids in effluent from the polishing pond, the TSS in the effluent samples was considered as algal solids. The difference between the BOD5 for non-filtered samples without inhibitor (nfBOD5-nATU) with the BOD5 for filtered samples containing inhibitor (fBOD5-ATU) indicates the influence of both algae and nitrification on the BOD5 results.

RESULTS AND DISCUSSION The results of monthly average of effluent quality parameters from the Miandoab polishing pond are given in Table 2. In addition to DO, pH, temperature and TSS, it also presents other parameters including BOD5 and COD, which were analyzed under various conditions. Alkaline pH and high DO in the effluent confirm the significant impact of algae on maturation ponds. These conditions for DO and pH also provide an appropriate environment for nitrifier organisms. Both algae and nitrification could affect effluent quality. The monthly mean variations of BOD5 in different modes are shown in Figure 2 for determining NBOD, algal BOD and the simultaneous effects of them on BOD5.

The quality parameters for polishing pond effluent shown in Table 2 and Figure 2 indicate that the annual mean TBOD5 of polishing pond effluent, considered as the nfBOD5-nATU term, was 61 mg/L; this consisted of 27.4 mg/L (44.92%) for algal CBOD5, 26.6 mg/L (43.61%) for TNBOD5 and 7 mg/L (11.47%) for sCBOD5. Considering the long retention time in lagoons and ponds, the sCBOD5 and non-algal solids in effluent from the system of aerated lagoons followed by polishing pond is less than 10 mg/L (Mara & Pearson ). In fact, a vast portion of the effluent’s TSS and CBOD5 is caused by algae growth (Gronlund et al. ). Conversely, due to the optimum growth conditions for nitrifiers, the TNBOD5 mainly comprises TBOD5 in the lagoon and pond effluent. Thereby, usage of TBOD5 to assess the effluent quality of secondary wastewater treatment processes, particularly stabilization ponds and lagoons, is regarded as an uncertain parameter. Instead, the CBOD5 for the filtered samples are used according to EPA guidelines (Rich ). In EU countries, the fCBOD5 and fCOD of lagoon effluents for discharge to surface waters are set at 25 mg/L and 125 mg/L, respectively; while TSS in non-filtered effluent from lagoons can reach up to 150 mg/L (Mara & Pearson ; Gronlund et al. ). Despite the improving efficiency of biological activity in summer, according to Figure 2, the average TBOD5 in polishing pond effluent decreased in warm weather due to optimum growth conditions for algae and nitrifying bacteria. Neglecting these findings may be responsible for falsely attributing the design and operation of wastewater treatment plant as defective. RMANOVA showed that the differences between the annual means BOD5 in the four groups discussed in material and methods were statistically significant (P < 0.001), and differences in the monthly means within these groups were also significant (P < 0.001).

98 H. Khorsandi et al.

The quality parameters of efﬂuent from the system of facultative partially mixed lagoons followed by polishing pond in the Miandoab WWTP, Iran

Month

July 2009 August 2009

DO mg/L

Temp. C

TSS mg/L

fCOD mg/L

nfCOD mg/L

fBOD5-NATUa

fBOD5-ATUb

nfBOD5-ATUc

nfBOD5-nATUd

1.2 ± 0.9

7.8 ± 0.1

25.2 ± 1.3

84.9 ± 22.6

64 ± 7.4

120.8 ± 24.2

18.8 ± 3.1

11.9 ± 1.9

58 ± 7.1

104.2 ± 29.2

4.1 ± 3.4

8 ± 0.1

24.4 ± 0.9

51.1 ± 22.6

63.2 ± 5.7

98 ± 17.5

9 ± 2.7

8.8 ± 2.6

33 ± 9.3

60.7 ± 21.4

September 2009

3.2 ± 1.5

8 ± 0.1

24.1 ± 0.9

45.6 ± 2.9

71.9 ± 4.6

86.5 ± 3.1

9.3 ± 2

8.3 ± 2

19.4 ± 2.1

October 2009

6.9 ± 1.7

8.2 ± 0.1

19 ± 1.3

57.1 ± 10.1

55.5 ± 9.7

89.4 ± 6.9

7.5 ± 3.8

5.7 ± 1.7

21.5 ± 7.8

26.3 ± 8.8

7.7 ± 1.4

8.2 ± 0.1

16.3 ± 1.1

60.6 ± 12.1

66.5 ± 3

100.7 ± 10.2

5.5 ± 0.9

4.9 ± 0.9

27.8 ± 4.8

41.3 ± 5.3

8.3 ± 0.1

10.4 ± 1.4

65.8 ± 8.7

110.4 ± 12.4

5.8 ± 2.8

4.7 ± 1.9

22.2 ± 3.7

31.7 ± 7.2

34 ± 12.1

November 2009

December 2009

January 2010

12.9 ± 1.2

8.4 ± 0.1

9.9 ± 1.3

87.8 ± 4.5

46 ± 6.2

132.3 ± 7.9

4.4 ± 0.5

3.6 ± 0.8

22.7 ± 1

57.7 ± 5

February 2010

11.7 ± 1.4

8.3 ± 0.1

8.8 ± 0.8

86 ± 3.9

36.8 ± 4.8

128.7 ± 4.8

5.2 ± 1.1

4.5 ± 0.9

27.4 ± 2.2

55.1 ± 11.5

March 2010

10.6 ± 0.9

8.3 ± 0.1

9.5 ± 0.7

91.2 ± 2.9

43.9 ± 5.6

143.7 ± 6.2

6.7 ± 0.7

5.8 ± 0.8

33.3 ± 4.5

50.2 ± 19

April 2010

8.6 ± 2.6

8.3 ± 0.2

13.6 ± 2.9

106.3 ± 13.8

50.7 ± 5.6

162.8 ± 18.2

6.3 ± 1.7

5.7 ± 1.7

36.9 ± 7.1

55.8 ± 5.9

12.1 ± 2

66 ± 9.7

May 2010

2.8 ± 1.6

7.9 ± 0.2

19.2 ± 2.5

87.6 ± 14.4

48.7 ± 6.1

154.4 ± 12.1

8.4 ± 1.2

8.4 ± 2.1

46.2 ± 8.4

90.7 ± 9.6

June 2010

5.1 ± 3.9

7.9 ± 0.1

23.8 ± 2.7

72.6 ± 10.5

46.7 ± 8.2

114.5 ± 27.1

12.8 ± 4.1

9.1 ± 1.3

49.1 ± 6.6

94.7 ± 3

120

7 ± 4.6

8.1 ± 0.3

17.2 ± 6.6

75.9 ± 21.4

55.9 ± 12.5

120.7 ± 27

9 ± 5.4

7 ± 3.2

Total

34.4 ± 14.4

Nitrogenous and algal oxygen demand in efﬂuent from lagoon

Table 2

61 ± 30.5

The ﬁltered sample BOD5 in the absence of nitriﬁcation inhibitor.

The ﬁltered sample BOD5 in the presence of nitriﬁcation inhibitor.

Water Science & Technology

The non-filtered sample BOD5 in the presence of nitrification inhibitor. d The non-filtered sample BOD5 in the absence of nitrification inhibitor.

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Figure 2

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Nitrogenous and algal oxygen demand in efﬂuent from lagoon

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The monthly average BOD5 variations of effluent from a system of facultative partially mixed lagoons followed by a polishing pond in the Miandoab WWTP, Iran. (a) Nitrification effect on BOD5; (b) algae effect on BOD5; (c) nitrification and algae effects on BOD5; (d) comparison of algae and nitrification on BOD5.

RMANOVA made a clear distinction between the annual mean COD of filtered samples (fCOD) and non-filtered samples (nfCOD), and overall significant difference of the monthly mean within these groups (P < 0.001). Taking into consideration Table 2, the annual mean fCOD in polishing pond effluent was 55.9 mg/L, but algal COD with an annual average of 64.8 mg/L increased the total annual average COD to 120.7 mg/L, making up 53.7% of the total COD. Therefore, nitrogenous and algal oxygen demand result in an increase in the effluent TBOD5 and COD. These results are consistent with the findings of Rich () about aerated

lagoon systems consisting of four cells in series in order to undertake partial polishing of effluent (1999). However, Rich () did not use statistical analysis for the results. Figure 3 presents the result obtained from the monthly mean variations of algal COD, algal CBOD5 and TSS of the effluent. Moreover, the ratio of the monthly mean of algal CBOD5 and algal COD to TSS are shown in this diagram. A non-linear regression having R squared of 0.77 between the algal COD and the TSS demonstrates that algae have a significant impact on effluent quality, particularly on COD. Although linear correlation was further confirmed between

100

Figure 3

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Nitrogenous and algal oxygen demand in efﬂuent from lagoon

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The monthly average algal COD, algal CBOD5 and TSS, and correlation between the algal COD and algal CBOD5 with TSS. (a) Monthly variations of algal COD/TSS and algal CBOD5/TSS; (b) monthly variations of algal COD, algal CBOD5 and TSS; (c) correlation between algal COD and TSS; (d) relationship between algal CBOD5 and TSS.

them with a coefficient of 0.87 using one-way ANOVA (P < 0.001), it was less preferable due to it being less consistent with expected algal COD and TSS near the origin of zero. According to Figure 3, the annual average ratios of (algal COD)/TSS and (algal CBOD5)/TSS were 0.80 and 0.37, respectively, which are consistent with Sperling () results. To demonstrate the role of algae in quality parameter assessment of stabilization pond effluent, Sperling () showed that the ratios of (algal COD)/TSS and (algal CBOD5)/TSS in the effluent from stabilization ponds were 1–1.5 and 0.3–0.4, respectively. Figure 3(d), having a high degree of scatter, indicates that algal CBOD5 has low correlation with TSS due to the dependence of algal CBOD on TSS concentration and also on the biodegradability of carbonaceous organic compounds present in different types of algae. The algal green color and microscopic observations confirmed the presence of algae in the effluent. In addition, the mean of chlorophyll A during the experiment was 345 μg/L. Based on Mara’s () findings, as all wastewater stabilization pond algae are composed of ∼1.5 per cent chlorophyll A by weight, and because it is easily measured, expression of algal biomass in terms of chlorophyll A is more meaningful. Accordingly, 23 mg/L algal biomass, as shown in this study, would generate the theoretical oxygen demand (Yang et al. ; Dalrymple et al. ) of 28.5 mg/L. This computational finding is consistent with the experimental algal CBOD5.

It should be noted that the nature of algal BOD5, COD and TSS is different from that of raw sewage because the oxygen produced by the algae in receiving water is higher than the oxygen consumed by other aquatic organisms (Mara ). However, in local standard procedures, there is no assessment methodology for nitrified effluent containing algae. Therefore, some managers consider algal growth in lagoons and ponds as undesirable (Meneses et al. ). As a result, they try to improve and upgrade the treatment plant. However, algae are useful in the uptake of nitrogen and phosphorus from wastewater and are also beneficial to agriculture (Naddafi et al. ; Machibya & Mwanuzi ; Mansouri et al. ), because they will increase organic content over time, and finally increase the water retention capacity of soil (Papadopoulos et al. ; Ensink et al. ; Sharafi et al. ).

CONCLUSION The aim of this study is to investigate the analysis of nitrogenous and algal oxygen demand in the efﬂuent from system of aerated lagoons followed by polishing pond. Findings demonstrated that the annual mean of total 5-day biochemical oxygen demand in the efﬂuent from a polishing pond consisted of 44.92% as algal CBOD, 43.61% as NBOD and 11.47% as soluble CBOD. The annual average ratios of (algal COD)/TSS and (algal CBOD5)/TSS were 0.80 and

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0.37, respectively. Based on this study, in addition to monitoring fCBOD5, fCOD, and TSS, other parameters such as the algal CBOD5, TNBOD5, and chlorophyll should be determined for assessment of the treatment plant’s performance and correct interpretation of the efﬂuent quality from a stabilization pond or lagoon.

ACKNOWLEDGEMENTS The authors would like to thank the Municipal Water and Wastewater Company of West Azerbaijan province, Iran, especially Mr Shamchi (CEO) and Mr Heydarlou (Deputy Director of Sewage) for their cooperation in conducting this research. We also gratefully acknowledge Mr J. Khorsandi for editing the text.

REFERENCES APHA, AWWA, WEF  Standard Methods for the Examination of Water and Wastewater. 21st edn, New York, USA. Barth, E. F.  To inhibit or not to inhibit: that is the question. J. Water Pollut. Control Fed. 53 (11), 1651–1652. Bitton, G.  Wastewater Microbiology. 4th edn, WileyBlackwell, New Jersey, USA. Chapman, K., James, H. & Muirhead, W.  Minimizing the impact of nitrification in the BOD5 test. Oper. Forum 8 (9), 1–11. Dague, R. E.  Inhibition of nitrogenous BOD and treatment plant performance evaluation. J. Water Pollut. Control Fed. 53 (12), 1738–1741. Dalrymple, O. K., Halfhide, T., Udom, I., Gilles, B., Wolan, J., Zhang, Q. & Ergas, S.  Wastewater use in algae production for generation of renewable resources: a review and preliminary results. Aquat. Biosyst. 9 (2), 1–11. Ensink, J. H. J., Van, D. H. W., Mara, D. D. & Cairncross, S.  Waste stabilization pond performance in Pakistan and its implications for wastewater use in agriculture. Urban Water J. 4 (4), 261–267. Gerardi, M. H.  Nitrification and Denitrification in the Activated Sludge Process. Wiley-Interscience, New York, USA. Griffiths, E. W.  Removal and Utilization of Wastewater Nutrients for Algae Biomass and Biofuels. MS Thesis. Utah State University, Biological and Irrigation Engineering Department. Gronlund, E., Johansson, E., Hanaes, J. & Falk, S.  Seasonal microalgae variation in a subarctic wastewater stabilization pond using chemical precipitation. VATTEN 60, 239–249. Hall, J. C. & Foxen, R. J.  Nitrification in the BOD test increases POTW noncompliance. J. Water Pollut. Control Fed. 55 (12), 461–1469.

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Khorsandi, H., Movahedyan, H., Bina, B. & Farrokhzadeh, H. a Innovative anaerobic/upflow sludge blanket filtration bioreactor for phosphorus removal from wastewater. Environ. Technol. 32 (5), 499–506. Khorsandi, H., Movahedyan, H., Bina, B. & Farrokhzadeh, H. b Innovative anaerobic/upflow sludge blanket filtration combined bioreactor for nitrogen removal from municipal wastewater. Int. J. Environ. Sci. Tech. 8 (2), 417–424. Machibya, M. & Mwanuzi, F.  Effect of low quality effluent from wastewater stabilization ponds to receiving bodies, case of Kilombero sugar ponds and Ruaha River, Tanzania. Int. J. Environ. Res. Public Health 3 (2), 209–216. Mansouri, B., Rezaaei, Z. & Mansouri, A.  Assessing wastewater stabilization ponds for agricultural use: a case study from Iran. Int. J. Current Res. 3 (6), 59–62. Mara, D.  Domestic Wastewater Treatment in Developing Countries. Earthscan, UK. Mara, D. & Pearson, H.  Design Manual for Waste Stabilization Ponds in Mediterranean Countries. Lagoon Technology International Ltd, Leeds, UK. Meneses, C. G., Saraiva, L. B., Melo, H. N., de Melo, J. L. & Pearson, H. W.  Variations in BOD, algal biomass and organic matter biodegradation constants in a wind-mixed tropical facultative waste stabilization pond. Water Sci. Technol. 51 (12), 183–190. Naddafi, K., Jaafarzadeh, N., Mokhtari, M., Zakizadeh, B. & Sakian, M. R.  Effects of wastewater stabilization pond effluent on agricultural crops. Int. J. Environ. Sci. Technol. 1 (4), 273–277. Papadopoulos, A., Parissopoulos, G., Papadopoulos, F., Papagianopoulou, A. & Karteris, A.  Impact of effluent recirculation on stabilization pond performance. Water, Air Soil Pollut.: Focus 4 (4–5), 157–167. Ramaraj, R., Tsai, D. D. W. & Chen, P. H.  Algae growth in natural water resources. J. Soil Water Conserv. 42 (4), 439–450. Rich, L. G.  High Performance Aerated Lagoon Systems. American Academy of Environmental Engineers, Annapolis, USA. Sharafi, K., Davil, M. F., Heidari, M., Almasi, A. & Taheri, H.  Comparison of conventional activated sludge system and stabilization pond in removal of chemical and biological parameters. Int. J. Environ. Health Eng. 1, 38. Sperling, M. V.  Waste Stabilization Ponds. IWA Publishing, London, UK. Stander, L. & Theodore, L.  Environmental Regulatory Calculation Handbook. Jon Wiley & Sons Inc., New Jersey, USA. Tchobanoglous, G., Burton, F. L. & Stensel, H. D.  Wastewater Engineering (Treatment and Reuse). 4th edn, McGraw-Hill Inc., New York, USA. Vesilind, P. A., Morgan, S. M. & Heine, L. G.  Introduction to Environmental Engineering. Cengage Learning, Stomford, US. Yang, X. E., Wu, X., Hao, H. L. & He, Z. L.  Mechanisms and assessment of water eutrophication. J. Zhejiang Univ. Sci. B 9 (3), 197–209.

First received 22 December 2013; accepted in revised form 8 April 2014. Available online 22 April 2014

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Removal of hazardous pharmaceutical dyes by adsorption onto papaya seeds Caroline Trevisan Weber, Gabriela Carvalho Collazzo, Marcio Antonio Mazutti, Edson Luiz Foletto and Guilherme Luiz Dotto

ABSTRACT Papaya (Carica papaya L.) seeds were used as adsorbent to remove toxic pharmaceutical dyes (tartrazine and amaranth) from aqueous solutions, in order to extend application range. The effects of pH, initial dye concentration, contact time and temperature were investigated. The kinetic data were evaluated by the pseudo ﬁrst-order, pseudo second-order and Elovich models. The equilibrium was evaluated by the Langmuir, Freundlich and Temkin isotherm models. It was found that adsorption favored a pH of 2.5, temperature of 298 K and equilibrium was attained at 180–200 min. The adsorption kinetics followed the pseudo second-order model, and the equilibrium was well

Caroline Trevisan Weber Gabriela Carvalho Collazzo Marcio Antonio Mazutti Edson Luiz Foletto Guilherme Luiz Dotto (corresponding author) Chemical Engineering Department, Federal University of Santa Maria, 97105-900, Santa Maria, Brazil E-mail: guilherme_dotto@yahoo.com.br

represented by the Langmuir model. The maximum adsorption capacities were 51.0 and 37.4 mg g 1 for tartrazine and amaranth, respectively. These results revealed that papaya seeds can be used as an alternative adsorbent to remove pharmaceutical dyes from aqueous solutions. Key words

| adsorption, amaranth, dyes, papaya seeds, tartrazine

INTRODUCTION Several industries, such as pharmaceuticals, textiles and tanneries generate dye containing effluents. The inadequate release of these effluents into the environment can cause adverse effects on the aquatic ecosystem and on human life (Gupta & Suhas ). Specifically, tartrazine and amaranth are toxic pharmaceutical azo dyes and can be degraded into aromatic amines, which are very harmful. These dyes are carcinogenic, mutagenic, cause hyperactivity in children, hypersensitivity and allergenic reactions (Kobylewski & Jacobson ). Accordingly, the search for alternatives to remove these dyes is necessary. Many techniques have been used to treat dye containing effluents (Gupta & Suhas ). Among these, adsorption with activated carbon is one of the most commonly employed methods for the removal of synthetic dyes from aqueous effluents, due its simplicity and high efficiency (Foletto et al. ; Zhang & Ou ). However, the costs of production and regeneration of activated carbon are high (Gupta & Suhas ). Therefore, lowcost materials from industrial or vegetable wastes may be used as promising adsorbents in order to make the doi: 10.2166/wst.2014.200

adsorption process less expensive (Weng et al. ; Yang et al. ). Some studies reported that papaya seeds were adequate for the removal of dyes from aqueous solutions. Papaya seeds were effective for removing methylene blue (Hameed ), from leather (Weber et al. ) and textile (Foletto et al. ) dyes. However, an adsorbent should be adequate for a wide range of dyes, and the use of papaya seeds for the adsorption of hazardous pharmaceutical dyes has not been reported yet. The use of papaya seeds as adsorbent is very interesting because it is a widely available and low-cost agricultural residue (Hameed ). In this context, the aim of this work was to investigate the use of papaya seeds as an alternative adsorbent for the removal of tartrazine and amaranth from aqueous solutions. First, the effect of pH was investigated. Next, kinetic studies were carried out under different initial dye concentrations and the pseudo ﬁrst-order, pseudo secondorder and Elovich models were ﬁtted to the experimental data. Finally, the equilibrium was evaluated under different temperatures by the Langmuir, Freundlich and Temkin models.

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Adsorption of pharmaceutical onto papaya seeds

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MATERIALS AND METHODS Adsorbent and dyes Papaya seeds were prepared and characterized according our previously published work (Foletto et al. ). Papaya fruits were collected at a local farm. The seeds were removed from the fruit, washed, oven dried at 85 C for 12 h and ground in a knife-mill (Wiley Mill Standard, 03, USA). The resulting material was sieved, and a portion with a particle diameter between 350 and 450 μm was used in the experiments. The seeds presented a speciﬁc surface area of 0.23 m2 g 1 and pore diameter of 332 Å. The scanning electron microscopy image showed papaya seeds containing a heterogeneous, irregular and rough surface. Tartrazine (molecular weight 534.4 g mol 1; color index 19140; and λmax ¼ 426 nm) and amaranth (molecular weight 604.5 g mol 1; color index 16185; and λmax ¼ 521 nm) with purity higher than 85% were supplied by a local manufacturer (Duas Rodas Ind.). The chemical structure of the dyes is presented in Figure 1. All solutions were prepared using distilled water and the reagents were of analytical grade. W

Figure 2

Effect of pH on the adsorption of pharmaceutical dyes onto papaya seeds (C0 ¼ 20 mg L

and T ¼ 298 K).

Adsorption assays The adsorption assays were carried out in a batch system using a thermostated orbital shaker (Tecnal, TE-141, Brazil). The effects of pH (2.5–8.5) (which were adjusted using H2SO4 and NaOH), initial dye concentration (20–70 mg L 1), contact time (0–300 min) and temperature (298–328 K) were investigated. These conditions were determined by preliminary tests. 0.05 g of adsorbent was added into 100 mL of dye aqueous solution in determined experimental conditions. The resulting solution was continuously stirred at 100 rpm until the equilibrium was reached. Aliquots were taken at various time intervals, centrifuged at 4,000 rpm for 20 min (Centribio, 80-2B, Brazil)

Figure 3

Kinetic curves for the adsorption of tartrazine (a) and amaranth (b) onto papaya seeds (pH 2.5 and 298 K).

Figure 1

Chemical structure of the dyes: (a) tartrazine and (b) amaranth.

and ﬁltered (Whatman No. 40) before analysis. The dye concentration was determined by spectrometry (Spectro Vision, T6-UV, Brazil). The mass of dye adsorbed per gram of

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adsorbent at any time (qt (mg g 1)) and at equilibrium (qe (mg g 1)) were calculated as follows (Crini & Badot ): qt ¼

ðC0 Ct Þ V W

(1)

qe ¼

ðC0 Ce Þ V W

(2)

where C0, Ct and Ce (mg L 1) are the dye concentrations at t ¼ 0, at any time and at equilibrium, respectively, W (g) is the adsorbent amount and V (L) is the volume of the solution. Kinetic and equilibrium models For both dyes, kinetic curves were obtained under different initial concentrations, ranging from 20 to 70 mg L 1. Pseudo ﬁrst-order (Lagergren ), pseudo second-order (Ho & Mckay ) and Elovich models (Zeldowitsch ) were ﬁtted to the experimental data in order to investigate the adsorption kinetics. These models are given by the following equations: qt ¼ q1 (1 exp ( k1 t))

Table 1

(3)

qt ¼

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t 1=k2 q22 þ ðt=q2 Þ

(4)

1 ln (1 þ abt) a

(5)

where k1 and k2 are the rate constants of pseudo ﬁrst-order and pseudo second-order models, respectively, in (min 1) and (g mg 1 min 1), q1 and q2 are the theoretical values for the adsorption capacity (mg g 1), a is the initial velocity due to dq/dt with qt ¼ 0 (mg g 1 min 1), b is the desorption constant of the Elovich model (g mg 1) and t is the time (min). The equilibrium curves were obtained at different temperatures (298, 308, 318 and 328 K). In order to ﬁt the experimental equilibrium data, Langmuir (Langmuir ), Freundlich (Freundlich ) and Temkin (Temkin ) isotherm models were applied. These models are given by the following equations: qe ¼

qm kL Ce 1 þ ðkL Ce Þ

(6)

qe ¼ kF Ce1=nF

(7)

Kinetic parameters for the adsorption of tartrazine and amaranth onto papaya seeds

Tartrazine Models

20 (mg L 1)

Amaranth 40 (mg L 1)

50 (mg L 1)

70 (mg L 1)

20 (mg L 1)

40 (mg L 1)

50 (mg L 1)

70 (mg L 1)

38.3

40.4

41.1

24.2

25.3

33.4

34.8

Pseudo ﬁrst-order q1 k1 R

ARE (%)

27.7 0.024

0.018

0.020

0.016

0.015

0.017

0.923

0.943

0.951

0.973

0.941

0.965

8.93

9.61

8.12

9.69

7.67

14.5

0.029

0.036

0.969

0.971

10.0

5.45

Pseudo second-order q2

27.8

40.0

41.9

45.4

21.2

22.1

31.5

32.8

0.0020

0.0027

0.0029

0.0030

0.0020

0.0022

0.0028

0.0030

0.993

0.995

0.991

0.998

0.999

0.995

0.997

ARE (%)

5.22

3.45

8.93

8.51

7.88

1.05

7.50

8.41

Elovich a b

0.160 20.66

0.140 20.07

0.138 20.87

0.134 20.01

0.196

0.193

7.67

8.02

0.958

0.971

0.965

0.980

0.961

ARE (%)

9.38

8.26

7.97

6.11

8.19

qe(exp)

33.3

42.8

44.8

46.0

28.0

0.956

0.160 11.34 0.952

10.15

10.03

31.6

35.3

0.163 12.54 0.978 7.58 35.6

105

qe ¼

C. T. Weber et al.

Adsorption of pharmaceutical onto papaya seeds

RT lnðAT Ce Þ B

Water Science & Technology

(8)

where qm is the maximum adsorption capacity (mg g 1), kL is the Langmuir constant (L mg 1), kF is the Freundlich constant (mg g 1) (mg L 1) 1/nF, 1/nF is the heterogeneity factor, R is the universal gas constant (kJ mol 1 K 1), T is the temperature (K), B is related to the adsorption heat and AT is the Temkin constant (L mg 1). The kinetic and isotherm parameters were found by non-linear regression using the software Statistica 6.0 (Statsoft, USA). The ﬁt quality was measured according to the coefﬁcient of determination (R2) and average relative error (ARE).

RESULTS AND DISCUSSION

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Later, the adsorption rate decreased considerably, with equilibrium being attained at 180–200 min. It was found that the qt values increased as a function of C0 increase (Figure 3). This behavior can be explained by two facts: (1) at higher values of C0, the concentration gradient between the bulk solution and the adsorbent external surface is higher, facilitating the external mass transfer; and (2) at higher values of C0, the mass flux of dye by surface diffusion is increased. Similar behavior was verified by Weber et al. () in the adsorption of Direct Black 38 onto papaya seeds. Table 1 shows the kinetic parameters for the adsorption of tartrazine and amaranth onto papaya seeds. The high values of the coefficient of determination (R 2 > 0.99) and the low values of the ARE (ARE < 10.0%) demonstrated that the pseudo second-order model was the more adequate to represent the adsorption kinetics of the pharmaceutical dyes onto papaya seeds. The values of q2 and k2 increased as a function of the C0 increase

Effect of pH The effect of pH was evaluated under the following conditions: initial dye concentration of 20 mg L 1, temperature of 298 K, 0.05 g of adsorbent and volume of solution 0.1 L. The results are shown in Figure 2. For both dyes, it can be seen (Figure 2) that the pH increase from 2.5 to 8.5 caused a strong decrease in adsorption capacity. This behavior can be explained on the basis of the point of zero charge of the papaya seeds (pHpzc ¼ 6.25) (Unuabonah et al. ). Under basic conditions (pH > pHpzc) the surface charge of papaya seeds is negative, leading to an electrostatic repulsion of the anionic dyes (see Figure 1). However, under acid conditions (pH < pHpzc) the hydrogen atoms (Hþ) in the solution tend to protonate the papaya seeds’ surface, facilitating electrostatic interactions with the anionic dyes. A similar trend was obtained by Rêgo et al. () in the adsorption of tartrazine and amaranth on chitosan ﬁlms. In this work, the more adequate pH for the adsorption of tartrazine and amaranth was 2.5, and so this pH was selected for the study sequence.

Kinetic study Kinetic curves were obtained at pH 2.5 and a temperature of 298 K. The effect of the initial dye concentration (20–70 mg L 1) was evaluated. The kinetic curves are shown in Figure 3. The kinetic curves revealed a fast adsorption process, where about 80–90% of saturation was attained at 80 min.

Figure 4

Equilibrium curves for the adsorption of tartrazine (a) and amaranth (b) onto papaya seeds (pH 2.5).

C. T. Weber et al.

106

Table 2

Adsorption of pharmaceutical onto papaya seeds

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Isotherm parameters for the adsorption of tartrazine and amaranth onto papaya seeds

Tartrazine Models

Amaranth

298 K

308 K

318 K

328 K

51.0

41.1

35.3

25.8

298 K

308 K

318 K

328 K

28.0

22.8

15.6

Langmuir qm

37.4

0.373

0.353

0.245

0.182

0.410

0.279

0.215

0.161

0.999

0.998

0.999

0.996

0.997

0.999

0.995

1.12

1.71

3.45

4.32

2.78

4.00

2.95

3.61

ARE (%) Freundlich kF

26.9

22.8

17.6

12.0

22.7

14.9

11.2

6.9

1/nF

0.144

0.149

0.162

0.169

0.119

0.148

0.161

0.178

0.951

0.943

0.948

0.963

0.970

0.972

0.941

0.958

ARE (%)

7.89

9.38

9.47

7.05

6.29

8.51

12.69

8.43

Temkin AT B R2

40.4 0.39 0.901

52.4 0.51 0.889

18.4

11.1

0.54

214.3

0.74

0.898

0.891

0.63 0.931

39.0 0.75 0.877

19.2 0.86 0.907

9.2 1.21 0.904

ARE (%)

11.28

10.18

13.40

13.21

10.12

15.07

11.50

12.11

qe(exp)

46.0

39.1

32.2

23.0

35.6

26.7

21.4

14.2

(Table 1). This shows that more dye was adsorbed and the process was faster at high concentrations.

Equilibrium study The equilibrium curves were obtained at 298, 308, 318 and 328 K, using 0.05 g of adsorbent at pH 2.5. The curves are shown in Figure 4. Figure 4 shows that the adsorption isotherm curves were characterized by an initial step with an increase in the adsorption capacity followed by a convex shape. The initial step indicates a great papaya seeds–dyes affinity and numerous readily accessible sites. The convex shape suggests the formation of a monomolecular layer of the dyes on the papaya seeds’ surface (Dotto et al. ). The adsorption capacity increased with temperature decrease reaching maximum values at 298 K (Figure 4). Similar behavior was verified by Dotto et al. () in the adsorption of Acid Blue 9 and FD&C red 40 on Spirulina platensis. The experimental equilibrium curves were fitted with Langmuir, Freundlich and Temkin models. These results are shown in Table 2. The higher values of the coefficient of determination (R 2 > 0.99) and the lower values of the ARE (ARE < 5.0%) were found for the Langmuir model. In this way, the

Langmuir model was the more appropriate to represent the adsorption of pharmaceutical dyes on papaya seeds. The kL values increased as a function of the temperature decrease, indicating that the papaya seeds–dyes affinity was higher at 298 K. A similar trend was verified for the qm parameter, being the maximum adsorption capacities 51.0 and 37.4 mg g 1, for tartrazine and amaranth, respectively, obtained at 298 K. Comparing these values with the literature cited in this work, it can be affirmed that papaya seeds are an alternative adsorbent to remove pharmaceutical dyes from aqueous solutions.

CONCLUSIONS Papaya seeds were tested as adsorbent to remove tartrazine and amaranth from aqueous solutions. The adsorption favored high concentrations, pH of 2.5, temperature of 298 K; equilibrium was attained at 180–200 min. The adsorption kinetics followed the pseudo second-order model and equilibrium was well represented by the Langmuir model. The maximum adsorption capacities were 51.0 and 37.4 mg g 1, for tartrazine and amaranth, respectively. These results revealed that papaya seeds can be used as an alternative adsorbent to remove pharmaceutical dyes from aqueous solutions.

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Adsorption of pharmaceutical onto papaya seeds

REFERENCES Crini, G. & Badot, P. M.  Application of chitosan, a natural aminopolysaccharide, for dye removal from aqueous solutions by adsorption processes using batch studies: a review of recent literature. Progress in Polymer Science 33 (4), 399–447. Dotto, G. L., Lima, E. C. & Pinto, L. A. A.  Biosorption of food dyes onto Spirulina platensis nanoparticles: equilibrium isotherm and thermodynamic analysis. Bioresource Technology 103 (1), 123–130. Foletto, E. L., Weber, C. T., Paz, D. S., Mazutti, M. A., Meili, L., Bassaco, M. M. & Collazzo, G. C.  Adsorption of leather dye onto activated carbon prepared from bottle gourd: equilibrium, kinetic and mechanism studies. Water Science and Technology 67 (1), 201–209. Foletto, E. L., Weber, C. T., Mazutti, M. A. & Bertuol, D. A.  Application of papaya seeds as a macro-/mesoporous biosorbent for the removal of large pollutant molecules from aqueous solution: equilibrium, kinetic and mechanism studies. Separation Science and Technology 48, 2817–2824. Freundlich, H.  Over the adsorption in solution. Journal of Physics Chemistry 57 (A), 358–471. Gupta, V. K. & Suhas  Application of low-cost adsorbents for dye removal: a review. Journal of Environmental Management 90 (8), 2313–2342. Hameed, B. H.  Evaluation of papaya seeds as a novel non-conventional low-cost adsorbent for removal of methylene blue. Journal of Hazardous Materials 162 (2–3), 939–944. Ho, Y. S. & Mckay, G.  Kinetic models for the sorption of dye from aqueous solution by wood. Process Safety and Environmental Protection 76 (B2), 183–191. Kobylewski, S. & Jacobson, M. F.  Food Dyes a Rainbow of Risks. Center for Science in the Public Interest, Washington, USA.

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Lagergren, S.  About the theory of so-called adsorption of soluble substances. Kungliga Svenska Vetenskapsakademiens 24 (4), 1–39. Langmuir, I.  The adsorption of gases on plane surfaces of glass, mica and platinum. Journal of the American Chemical Society 40 (9), 1361–1403. Rêgo, T. V., Cadaval, T. R. S., Dotto, G. L. & Pinto, L. A. A.  Statistical optimization, interaction analysis and desorption studies for the azo dyes adsorption onto chitosan ﬁlms. Journal of Colloid and Interface Science 411 (1), 27–33. Temkin, M. I.  Adsorption equilibrium and the kinetics of processes on non-homogeneous surfaces and in the interaction between adsorbed molecules. Journal of Physical Chemistry URSS 15 (3), 296–332. Unuabonah, E. I., Adie, G. U., Onah, L. O. & Adeyemi, O. G.  Multistage optimization of the adsorption of methylene blue dye onto defatted Carica papaya seeds. Chemical Engineering Journal 155 (3), 567–579. Weber, C. T., Foletto, E. L. & Meili, L.  Removal of tannery dye from aqueous solution using papaya seed as an efﬁcient natural biosorbent. Water, Air and Soil Pollution 224 (1), 1427–1438. Weng, C. H., Chang, E. E. & Chiang, P. C.  Characteristics of new coccine dye adsorption onto digested sludge particulates. Water Science and Technology 44 (10), 279–284. Yang, F., Song, X. & Yan, L.  Preparation of cationic waste paper and its application in poisonous dye removal. Water Science and Technology 67 (11), 2560–2567. Zeldowitsch, J.  Über den mechanismus der katalytischen oxydation von CO an MnO2. Acta Physicochemical URSS 1 (3–4), 449–464. Zhang, J. X. & Ou, L. L.  Kinetic, isotherm and thermodynamic studies of the adsorption of crystal violet by activated carbon from peanut shells. Water Science and Technology 67 (4), 737–744.

First received 8 October 2013; accepted in revised form 10 April 2014. Available online 23 April 2014

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Study of constructed wetlands efﬂuent disinfected with ozone N. D. Miranda, E. L. Oliveira and G. H. R. Silva

ABSTRACT The purpose of this research was to study the disinfection of sanitary effluent from constructed wetlands, evaluating the oxidation of organic matter, the formation of formaldehyde, as well as the efficiency of total coliforms and Escherichia coli inactivation. A constant flow of ozone was applied to the batch system in 5 and 10 mg.O3 L 1 doses with contact times of 5 and 10 min. This study revealed that the average values of formaldehyde formation ranged between 259.00 and 379.00 μg L 1, which means that the values are within World Health Organization recommended values. The total coliforms and E. coli showed complete inactivation in almost all tests. The dose of ozone 5 mg.O3 L 1 and contact

N. D. Miranda (corresponding author) E. L. Oliveira G. H. R. Silva Faculdade de Engenharia Campus de Bauru, Departamento de Engenharia Civil, UNESP – Univ. Estadual Paulista, Av. Eng. Luiz Edmundo Carrijo Coube, 14-01 Vargem Limpa, CEP 17033-360, Bauru – SP, Brazil E-mail: nathaliedmiranda@yahoo.com.br

time of 5 min were sufficient for a significant reduction of the concentration levels of pathogens in constructed wetlands effluent with similar characteristics, thus allowing for its agricultural reuse. Key words

| by-products, disinfection, sanitary wastewater, ozone

INTRODUCTION Constructed wetlands are controlled systems that simulate and accelerate the conditions found in natural wetlands. Their use as a sanitary effluent treatment system has intensified in recent decades, especially in small wastewater treatment plants (WWTPs) (Cooper ). Accordingly, it has been highlighted as a low-cost treatment for both construction and operation, the construction of which can be made near the site where the effluent is generated and can be operated without specific training (Campos et al. ). Ghermandi et al. () reported 1–2 log-removal for a range of indicator pathogens. However, when the aim is for reuse of effluent, efficient disinfection is required for ensuring both public and environmental health parameters (Domenjoud et al. ). Most of the WWTPs in Brazil do not have any specific steps for disinfection. The alternatives most commonly used in Brazil are maturation ponds and chlorination. Studies conducted on chlorine, however, demonstrate that this technology is not effective in eliminating some pathogenic agents and this can enable the formation of trihalomethanes (THMs) (White ); moreover, maturation ponds require a large area for implementation. An alternative disinfectant, ozone, is used in this study. Ozone has proved to be very effective, especially in the inactivation of coliform bacteria (Costa & Daniel ). In doi: 10.2166/wst.2014.202

addition, ozone may also increase the concentration of oxygen in the water, reduce organic matter, and remove color (Silva et al. ). Regarding the generation of by-products, the nature of the reaction mechanism of ozone with organic matter (direct or indirect reaction) determines the products formed. The by-products produced during ozonation include speciﬁc compounds such as aldehydes, ketones, and carboxylic acids as the major by-products (Dabrowska et al. ; Huang et al. ; Karnik et al. ). Some of these by-products are of particular concern due to their mutagenicity and carcinogenicity (Karnik et al. ). Aldehydes are a major by-product of ozonation and are of interest due to their effects on health. The major aldehydes formed are formaldehyde, acetaldehyde, glyoxal, and methylglyoxal (Dabrowska et al. ; Karnik et al. ). Studies by Wert et al. () have shown that the formaldehyde formed is an aldehyde, which is formed in higher concentrations during ozonation. Therefore, all treatment processes must be optimized to control such reactions; concentrations will be a problem to be resolved during practical usage (Mehrjouei et al. ). This research evaluated the disinfection of sanitary efﬂuent from a constructed wetland using ozone, taking into

109

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Study of constructed wetlands efﬂuent disinfected with ozone

consideration the results of the efﬁciency of total coliforms and Escherichia coli inactivation, organic matter oxidation, and formaldehyde formation.

Water Science & Technology

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2014

did not affect the results, because the off-gas and the final dissolved residual ozone did not show a significant difference. The assays were performed in triplicate for each dose and contact time, and 2 L samples of ozonated effluent were taken to analyze the performance of ozonation assays.

MATERIALS AND METHODS Analytical methods Effluent source Effluent samples for the constructed wetland were taken from the WTP at the State University of São Paulo ‘Júlio de Mesquita Filho’ (UNESP), Bauru Campus, Sao Paulo, Brazil. The constructed wetland is composed of five stages of treatment: preliminary treatment comprising screening, preliminary clarifier, and a hydrodynamic static sieve; secondary treatment comprising three beds of horizontal subsurface flow constructed wetlands (HSFCW – 9.0 × 4.5 m); and tertiary treatment comprising three beds of vertical subsurface flow constructed wetlands (VSFCW – 3.0 × 3.0 m). The plant species used is vetiver grass (Chrysopogon zizanioides), and the hydraulic load was 58 L m 2 d 1. The samples were taken only at the end of the treatment. A volume of 50 L of effluent was taken, in the morning after the treatment, once a week, over 3 weeks. All ozonation assays were carried out on the same day that effluent was taken from constructed wetlands. For each ozonation assay, 11.94 L of effluent (1.28 m in ozonation column) was used. The average values of physicochemical characteristics of the constructed wetlands effluent were: pH 6.7; dissolved oxygen 2.96 mg L 1; turbidity 40.0 nephelometric turbidity units (NTU); true color 97 Pt-Co; chemical oxygen demand (COD) 36 mg L 1; total coliform 1.4 × 106, and E. coli 3.4 × 105 MPN/100 mL. Ozonation experiments The ozonation experiments were performed at batch-scale consisting of: an oxygen concentrator (EverFlo Respironics, Germany); an ozone gas generator (BRO3-Brasil Ozônio, Brazil); an ozonation column (diameter of 0.10 m and height of 1.70 m) containing all the inert materials; a plastic micro porous bubble diffuser (pore apertures of 10 μm) placed at the bottom of the column; and absorption bottles with potassium iodide for the destruction of ozone off-gas. Ozone was applied to the system in batches, in doses of 5 and 10 mg L 1 with contact times of 5 and 10 min for each dosage. The gas flow rates ranged between 0.47 and 1.49 L min 1, regulated by a rotameter, which is part of the aforementioned oxygen concentrator. However, this difference

The ozone concentration in the efﬂuent gas was measured according the iodometric method Cl 4500 B (Rice et al. ). The dissolved residual ozone was determined according the N,N-diethyl-p-phenylenediamine (DPD) method (German standard DIN 38 408). Parameters such as a COD (5220-D method), pH (4500Hþ B method), turbidity (method 2130 B), true color (method C 2120) total coliforms, E. coli (9221 method) (Rice et al. ), and formaldehyde (HACH-MBTH 8110 method) were analyzed. The signiﬁcant differences in the average values of the analyzed parameters were evaluated according to the statistical method proposed by Miller & Miller ().

RESULTS AND DISCUSSION Ozonation The average mass transfer efficiency was: 94.4% for 5 mg.O3 L 1 and contact times of 5 and 10 min; 90.4% for 10 mg.O3 L 1 and a contact time of 5 min; and 82.8% for doses of 10 mg/L and a contact time of 10 min (Table 1). Silva et al. () obtained a transfer efficiency of 76 and 79% for a dose of 5 mg.O3 L 1 and a contact time of 5 and 10 min, respectively. For 10 mg.O3 L 1 and contact times of 5 and 10 min, the efficiency was 71%. In their study, the temperature ranged between 20 and 30 C; in the present study the applied temperature ranged between 17 and 24 C. It is important to remember that the water temperature rise decreases the solubility of ozone in it; in other words, it influences its reaction rate (Gottschalk et al. ). The difference in transfer efficiency between Silva et al. () and this study may be due to temperature. There was a significant difference (P ¼ 0.05) in the average values of consumed ozone only between ozone doses. In addition, there was not a significant difference (P ¼ 0.05) in the average values of the temperature applied. Therefore, the temperature did not affect the ozone consumed at the different doses and contact times applied in this study. There were increases in the amounts of dissolved oxygen in the ozonated effluent (Table 1), and the high W

N. D. Miranda et al.

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Table 1

Study of constructed wetlands efﬂuent disinfected with ozone

Water Science & Technology

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Average concentrations of consumed ozone, dissolved residual ozone; off-gas; efficiencies in the mass transference of ozone and concentrations of dissolved oxygen for the effluent from the constructed wetlands and ozonated effluent for dosages of 5 and 10 mg.O3 L 1 and contact times of 5 and 10 min Dissolved oxygen Contact time (min)

Ozone applied (mg.O3 L 1)

Off-gas (mg.O3 L 1)

Final dissolved residual ozone (mg.O3 L 1)

Consumed ozone (mg.O3 L 1)

Efﬂuent constructed wetlands (mg L 1)

Ozonated efﬂuent (mg L 1)

0.19

0.09

4.72

2.96

18.26

0.19

0.09

4.72

2.96

17.90

0.75

0.21

9.04

2.96

17.75

1.59

0.13

8.28

2.96

17.71

values of dissolved oxygen found (above saturation value) could be due to it being measured immediately after the ozonation assays. Turbidity There was no significant difference between the doses and contact times applied for the removal of turbidity, which has averaged 24.4% (Table 2). A significant removal of turbidity was not observed during this study, due to its good effluent quality in this aspect and because the constructed wetlands had already promoted efficient removal of any turbidity, showing mean values of 10.79 NTU.

Table 2

True color Color plays a significant role in the elementary assessment of the quality of water or wastewater, especially for a first print (Mehrjouei et al. ). The removal of true color was on average 54% (Table 2); however, there was no significant difference between the doses and the contact times. This means its appearance is not unpleasant to the public eye. pH The average constructed wetland effluent pH values were 6.7; for the ozonated effluent, the minimum and the maximum values were, respectively, 6.5 and 6.7 (Table 3).

Average values of turbidity and true color for the efﬂuent from the constructed wetlands and ozonated efﬂuent for doses of 5 and 10 mg.O3 L 1 and contact times of 5 and 10 min True color (Pt-Co)

Turbidity (NTU)

Ozone applied

Contact

Consumed ozone

(mg.O3 L 1)

time (min)

(mg.O3 L 1)

Efﬂuent constructed wetlands

Ozonized efﬂuent

Efﬂuent constructed wetlands

Ozonized efﬂuent

4.72

97 ± 20

54 ± 16

10.79 ± 4.56

8.76 ± 4.55

4.72

97 ± 20

41 ± 29

10.79 ± 4.56

7.01 ± 3.90

9.04

97 ± 20

39 ± 34

10.79 ± 4.56

8.19 ± 3.87

8.28

97 ± 20

43 ± 18

10.79 ± 4.56

8.65 ± 4.17

Table 3

Average values of ambient temperature, gas flow rate, pH and COD for the effluent from the constructed wetlands and ozonated effluent and values of formaldehyde formation for doses of 5 and 10 mg.O3 L 1 and contact times of 5 and 10 min COD (mg L 1)

Ozone applied (mg.O3

Contact

Consumed ozone

L 1)

time (min)

(mg.O3 L 1)

Gas ﬂow rate

Efﬂuent constructed

Ozonated

Efﬂuent constructed

Ozonated

Formaldehyde

T ( C)

(L min 1)

wetlands

efﬂuent

wetlands

efﬂuent

formation (μg L 1)

4.72

19.8

1.49

36 ± 20

33 ± 20

6.7 ± 0.2

259.00

4.72

20.1

0.47

36 ± 20

28 ± 19

6.7 ± 0.2

6.6 ± 0.2

322.33

9.04

20.2

1.26

36 ± 20

28 ± 23

6.7 ± 0.2

6.5 ± 0.2

357.00

8.28

20.3

1.49

36 ± 20

29 ± 22

6.7 ± 0.2

379.00

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Study of constructed wetlands efﬂuent disinfected with ozone

The pH value is the most important parameter because it affects the process in two ways: it affects the stabilization of ozone; and it determines the degree of ionization of organic compounds (Boncz ). Regarding the interactions between ozone and organic matter, the pH determines the predominant mechanism. Oxidation occurs directly via acid pH (molecular ozone) and indirectly via basic pH (hydroxyl radicals) and depending on this, both may occur simultaneously (Gottschalk et al. ). There was no significant change in average pH values after ozonation, regardless of time and dose applied. Therefore, as the pH did not change significantly among the tests, it can be concluded that pH was not responsible for the different results obtained by comparing each dose and contact time applied. Chemical oxygen demand The average values of COD in the ozonated and constructed wetlands effluent, and in the different doses and contact times applied, are shown in Table 3. The differences in the removal amounts were not significant (P ¼ 0.05) regardless of contact and dose applied. The mean reductions in COD were: 8.8 and 21.7% for applied ozone of 5 mg.O3 L 1 and contact times of 5 and 10 min, respectively; and were 21.7 and 18.9% for applied ozone of 10 mg.O3 L 1 and contact times of 5 to 10 min, respectively. Assirati (), in her study about ozonation of effluents from anaerobic ponds and sand filters, also confirmed no significant changes to the COD, indicating that the disinfectant did not cause significant chemical degradation of organic matter. Silva et al. () obtained a removal of 27–44% in similar applied dosages. However, the concentration of organic matter was greater than that of the effluent used in this study; values higher than 100 mg COD L 1 were reached. This may mean that the percentage of recalcitrant organic matter from their effluent was lower. Formaldehyde The average values for the formation of formaldehyde for the different contact times and doses applied are presented in Table 3. All dosages and contact times applied promoted the formation of formaldehyde in the ozonated effluent. There was only a significant difference between the dosage of 5 mg L 1 for 5 min and the dosage of 10 mg L 1 for contact times of both 5 and 10 min.

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As there was practically no variation in pH among doses and contact times applied, it may be concluded that it was not responsible for the different values of formaldehyde formation found. Another parameter, in addition to the pH and ozone dose, which can influence the formation of formaldehyde, according to Nawrocki et al. (), is the contact time. In their study, the contact time is of minor importance compared to the ozone dose for the potential of the aldehydes’ formation during the ozonation step. Although the contact time represents a minor importance in the formation of aldehydes, the results found in this study indicate that it was responsible for the difference in the formation of formaldehyde for the same dose applied. The results are consistent with Silva et al. (); in their data the formation of aldehydes varied according to the applied dose, which confirms that the concentration of formaldehyde is strongly dependent on the dose of ozone applied. According to Nawrocki & Kalkowska (), the formation of formaldehyde is formed by the direct route of ozonation, in other words, in an acidic environment; therefore, the availability of molecular ozone is the limiting factor in the production of aldehydes. The molecular ozone can interact with organic compounds through the Criegee mechanism (rupture of double bonds), electrophilic or nucleophilic reactions. The increase in pH benefits the decomposition of ozone, forming hydroxyl radicals, which will interact with organic matter (Langlais et al. ). In Brazil, like many other countries, there are no permissible levels for discharges of treated wastewater with formaldehyde into receiving water bodies. In order to make a comparison, some guidelines values were found for permissible levels for formaldehyde in water. In Japan the limit is 80 μg L 1 (Sugaya et al. ), in Poland the limit is 50 μg L 1 (Nawrocki & Biłozor ), for the World Health Organization (WHO ) it is 900 μg L 1, and in Australia it is 500 μg L 1 (Australian Drinking Water Guidelines ). The amounts of formaldehyde formation by ozonation were below the limits set by WHO and that of the Australian Drinking Water Guidelines but were above the limits of countries such as Japan and Poland. These values indicate that there should be some concern about the formation of formaldehyde because research (for example, Craftstorm et al. ; Scheuplein ; Tomkins et al. ) shows it has carcinogenic and mutagenic potential. Total coliforms and E. coli Total coliforms and E. coli were completely inactivated in almost all the tests conducted (Table 4). As the pH was

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Average inactivation of total coliforms and E. coli as a function of contact time and effective ozone concentrations consumed by ozonated efﬂuent Efﬂuent constructed wetlandsa

Ozone applied (mg.O3 L 1)

Contact time (min)

Ozonated efﬂuenta

Consumed ozone (mg.O3 L 1)

Total coliform

E. coli

4.72

1.4 × 106

3.4 × 105

log10 inactivation

Total coliform

E. coli

Total coliform

E. coli

8.24 × 101

5.43

4.23

4.99

4.72

1.4 × 10

6.15

5.53

9.04

1.4 × 106

3.4 × 105

9.28

5.38

5.53

8.28

6.15

5.53

10 a

1.4 × 10

3.4 × 10 3.4 × 10

Total coliform and E. coli in MPN/100 mL.

maintained between 6.5 and 6.7, there was no influence from pH on the total coliforms and E. coli inactivation since ozone disinfection has little dependence on the pH range of domestic effluents, which is between 6.0 and 8.0 (Silva et al. ). There were no significant differences (P ¼ 0.05) in removal between the different doses and contact times applied. This means it is possible to obtain high efficiency using small doses and contact times; in this case, 5 mg.O3 L 1 and 5 min, when the constructed wetlands effluent contains characteristics similar to those used in the present study, enabling both financial and time savings. In the study, by Paraskeva & Graham (), Silva et al. (), and Assirati (), who disinfected secondary and tertiary effluents, the contact time was not significant in the inactivation of total coliforms and E. coli, which also suggests that a lower dosage can achieve almost the same results. Nevertheless, in research by Assirati (), ozone doses between 8 and 21 mg.O3 L 1 were necessary to match the effluent from anaerobic ponds with those standards established by WHO () for agricultural reuse. Effluent from sand filters reached WHO’s standards with doses of up to 4 mg.O3 L 1. In the present study a dose of only 5 mg.O3 L 1 was necessary, with a contact time of 5 min, in order to cause significant reductions in the concentration of pathogens in wastewater constructed wetlands suitable for agricultural reuse (WHO ).

CONCLUSION According to this research, the ozonation obtained mass transfer efﬁciency above 80%. The pH did not present any substantial changes after ozonation. Consequently, the removal of coliform and E. coli, as well as the other parameters, was not affected. The percentage removal of COD was low due to the constructed wetlands efﬂuent already being of good

quality; furthermore, the average values of formaldehyde formation are within WHO () recommended values for water. Therefore, a dose of 5 mg.O3 L 1 with a contact time of 5 min is sufﬁcient to cause signiﬁcant reductions in the concentration of pathogens in wastewater constructed wetlands with similar characteristics, if it is intended for agricultural reuse (WHO ).

ACKNOWLEDGEMENTS The authors gratefully acknowledge the master’s scholarship Grant 2012/05297-9 and Research Project 2011/10816-2, São Paulo Research Foundation (FAPESP), UNESP – Univ. Estadual Paulista, the postgraduate program in Civil and Environmental Engineering.

REFERENCES Assirati, D. M.  Disinfection of Municipal Wastewater with Ozone for Agricultural Use. Master diss., The State University of Campinas (UNICAMP). Australian Drinking Water Guidelines 6, 2011/Australian Government National Health and Medical Research Council, Natural Resource Management Ministerial Council. Boncz, M. Á.  Selective Oxidation of Organic Compounds in Waste Water: By Ozone-based Oxidation Processes. Thesis, Wageningen University, Wageningen. Campos, J. C., Ferreira, J. A., Mannarino, C. F., Silva, H. R. & Borba, S. M. P.  Treatment of leachate from the landﬁll Pirai (RJ) using wetlands. Paper presented at the Sixth Symposium Italo Brazilian Sanitary and Environmental Engineering, Brazilian Association of Sanitary and Environmental Engineering, Vitória-ES. Cooper, P.  CWs after 25 years of application: A review of the developments that we have made and the problems that we still have to overcome. Paper presented at the 12th International Conference on Wetland Systems for Water Pollution Control. Venice, Italy, 4–9 October 2010.

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Costa, H. & Daniel, L.  Study of the behavior of ozone to disinfect household wastewater. PROSAB Call 3, Theme 2 – Disinfection of Sanitary Effluent, Removal of Pathogens and Hazardous Substances. Craftstorm, R. C., Fornace, A. J. Jr., Autrup, H., Lechner, J. F. & Harris, C. C.  Formaldehyde damage to DNA and inhibition of DNA repair in human bronchial cells. Science 220 (4593), 216–218. Dabrowska, A., Swietlik, J. & Nawrocki, J.  Formation of aldehydes upon ClO2 disinfection. Water Res. 37, 1161–1169. Domenjoud, B., Tatari, C., Esplugas, S. & Baig, S.  Ozonebased processes applied to municipal secondary effluents. Ozone Sci. Eng. 33 (3), 243–249. Ghermandi, A., Bixio, D., Traverso, P., Cersosimo, I. & Thoeye, C.  The removal of pathogens in surface-flow constructed wetlands and its implications for water reuse. Water Sci. Technol. 56 (3), 207–216. Gottschalk, C., Libra, J. A. & Saupe, A.  Ozonation of Water and Wastewater. Wiley-VCH, Weinheim, Germany. Huang, W. J., Fang, G. C. & Wang, C. C.  The determination and fate of disinfection by-products from ozonation of polluted raw water. Sci. Total Environ. 345, 261–272. Karnik, B. S., Davies, S. H., Baumann, M. J. & Masten, S. J.  The effects of combined ozonation and filtration on disinfection by-product formation. Water Res. 39, 2839–2850. Langlais, B., Reckhow, D. A. & Brink, D. R.  Ozone in Water Treatment: Application and Engineering. Lewis Publishers, CRC Press, Florida, USA, p. 569. Mehrjouei, M., Muller, S., Sekiguchi, K. & Moller, D.  Decolorization of wastewater produced in a pyrolysis process by ozone: enhancing the performance of ozonation. Ozone Sci. Eng. 32, 349–354. Miller, J. C. & Miller, J. N.  Statistic for Analytical Chemistry, 3rd edn, Ellis Horwood PTR Prentice Hal, New York, USA, p. 233. Nawrocki, J. & Biłozor, S.  Brominated oxidation by-products in drinking water treatment. J. Water SRT – Aqua 46, 304. Nawrocki, J. & Kalkowska, I.  Effect of pH and hydrogen peroxide on aldehyde formation in the ozonation process. J. Water SRT – Aqua 47, 50–56.

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Nawrocki, J., Swietlik, J., Raczyk-Stanislawiak, U., Dabrowska, A., Bilozor, S. & Ilecki, W.  Influence of ozonation conditions on aldehyde and carboxylic acid formation. Ozone: Science & Engineering 25 (1), 53–62. Paraskeva, P. & Graham, N. J. D.  Treatment of a secondary municipal effluent by ozone, UV and microfiltration: microbial reduction and effect on effluent quality. Desalination 186, 47–56. Rice, E. W., Baird, R. B., Eaton, A. D. & Clesceri, L. S. (eds)  Standard Methods for the Examination of Water and Wastewater, 22nd edn, American Public Health Association, American Water Works Association, Water Environment Federation, Washington, DC, USA. Scheuplein, R. J.  Formaldehyde: The Food and Drug Administration's Perspective. In: Formaldehyde – Analytical Chemistry and Toxicology (V. Turoski, ed.), American Chemical Society, Washington, DC. pp. 237–245 (Advances in Chemistry Series 210). Silva, G. H. R., Daniel, L. A., Bruning, H. & Rulkens, W. H.  Anaerobic effluent disinfection using ozone: Byproducts formation. Biores. Technology 101, 6992–6997. Sugaya, N., Nakagawa, T., Sakurai, K., Morita, M. & Onodera, S.  Analysis of aldehydes in water by head space-GC/MS. J. Health Sci. 47, 21–27. Tomkins, B., McMahon, J. & Caldwell, W.  Liquid chromatographic determination of total formaldehyde in drinking water. J. Assoc. Off. Analyt. Chem. 72, 835–839. Wert, E. C., Rosario-Ortiz, F. L., Drury, D. D. & Snyder, S. A.  Formation of oxidation byproducts from ozonation of wastewater. Water Res. 41 (7), 1481–1490. White, G. C.  Handbook of Chlorination and Alternative Disinfectants, 4th edn, Wiley-Interscience, John Wiley & Sons Inc., NY. WHO  Concise International Chemical Assessment Document (CICAD) no. 40. Formaldehyde. World Health Organization, Geneva. WHO  World Health Organization Guidelines for the Safe Use of Wastewater, Excreta and Greywater. Vol. 1: Policy and regulatory aspects. World Health Organization, Geneva.

First received 9 December 2013; accepted in revised form 11 April 2014. Available online 25 April 2014

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Comparison of sludge treatment by O3 and O3/H2O2 Zhao Yuxin, Wang Liang, Yu Helong, Jiang Baojun and Jiang Jinming

ABSTRACT This work focuses on the comparison of sludge decomposition caused by ozone (O3) alone and by ozone/hydrogen peroxide (O3/H2O2). The content of carbonaceous organic materials, nitrogenous compounds and phosphoric substances in sludge supernatant were measured. The release of soluble chemical oxygen demand, total nitrogen (TN) and total phosphorus (TP) caused by O3/H2O2 treatment were more than by O3 alone. As a result, it can be concluded that the efficiency of sludge breakup in O3/H2O2 was better than that in O3 alone. However, a peak appeared in both systems for the biodegradable substances such as carbohydrate. Carbohydrate could be used as the carbon source for denitrification, and the releasing of TN and TP may become an additional burden for a subsequent biological system. So, it was of benefit for the enhancement of cryptic growth and cost reduction by raising and maintaining the content of biodegradable substance and reducing the concentrations of the nitrogenous and phosphoric substances as far as possible. Therefore, sludge treated by O3/H2O2 with lower O3 dose would be more suitable than O3 alone. Key words

| cryptic growth, ozone, ozone/hydrogen peroxide, sludge reduction

Zhao Yuxin (corresponding author) Wang Liang Jiang Baojun Jiang Jinming Key Laboratory of Songliao Aquatic Environment, Ministry of Education, Jilin Jianzhu University, Changchun, 130118, Jilin Province, China E-mail: hitzyx@163.com Zhao Yuxin School of Municipal and Environmental Engineering, Harbin Institute of Technology, Harbin, 150090, Heilongjiang Province, China Yu Helong College of Information and Technology, Jilin Agricultural University, Changchun, 130118, Jilin Province, China

INTRODUCTION As an essential step for wastewater treatment, sludge disposal would account for 25–65% of the total plant operation cost (Liu ). Although traditional sludge treatment methods, such as landfill, agricultural application, sending to sea and incineration, are widely adopted, they are facing more and more environmental pressure and economic challenge with the increasing stringent environmental legislation as well as the reduction of land and ocean day by day (Ludovico et al. ). So, there is an urgent need to develop new and viable technologies to reduce the production of sludge and thus alleviate the sludge treatment load. Recently, sludge reduction by enhancing its cryptic growth has attracted extensive attention (Lan et al. ). In this process, sludge is dissolved first and then the solution would be degraded by microbes again when it returns to the activated sludge system. This way, the minimization of excess sludge production could be achieved by the enhancement of cell decay rate. The dissolution of sludge cells is the key step in cryptic growth, and thus many methods have been developed, such as biological treatment with thermophilic bacteria, mechanical disintegration, chemical treatment, thermal hydrolysis, ultrasound and the combination of them, in order to decompose the sludge flocs doi: 10.2166/wst.2014.185

and rigid cell wall of microorganisms (Zhao et al. ; Salsabil et al. ; Labelle et al. ; Tang et al. ; Ma et al. ). Among these methods, sludge ozonation has been considered as the most promising technology for its simple operation, high efficiency and no release of secondary pollutants (Chu et al. a). As a strong oxidant, ozone (O3) can destroy the sludge floc structure effectively and thus make the intracellular and extracellular release, where the reaction mechanism often includes direct oxidation by O3 and indirect oxidation by hydroxyl radical induced by O3 in alkaline solutions. The hydroxyl radical is a much more effective oxidant than O3, which can oxidize almost all pollutants with high reaction rates (Hoigne & Bader ). The generation of hydroxyl radical would be accelerated by adding some hydrogen peroxide (H2O2) to the ozone system, known as one effective advanced oxidation process for water and wastewater treatment (Natasa et al. ). But studies on sludge treatment by combined O3/H2O2 are limited. We previously examined the effect of O3/H2O2 treatment on sludge and found that the maximum efficiency of sludge disintegration can be achieved when the molar ratio of H2O2:O3 was 1.2 (Zhao et al. ). Although both O3 and O3/H2O2 can

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disintegrate sludge and change its properties, it is still unclear whether the sludge treated by O3/H2O2 is better for the cryptic growth than that treated by O3 alone. With the breakup of sludge ﬂocs and cells, a large amount of intracellular and extracellular substances are released into the surfactants, such as proteins, carbohydrates, nucleic acids, lipids and humic products (Kiyomi et al. ). When they return to the subsequent biological treatment system, the inﬂuent load of carbon, nitrogen and phosphorus would change, which is critical for the stable operation of biological treatment system. Therefore, the objective of this study is to compare the properties of sludge treated by combined O3/H2O2 and by O3 alone, based on the concentrations of carbonaceous organic material, nitrogenous compounds and phosphoric substances in the treated sludge supernatant.

MATERIAL AND METHODS Sludge treatment by ozone The sludge samples had been collected from the secondary sedimentation tank of Xijiao wastewater treatment plant in Changchun city (Jilin Province, China) and had been stored at 4 C for 2 days before use. The sludge has water content of 95% and pH of 7.03. The initial suspended solids concentrations (TS) were 8,000–10,000 mg/L and the volatile solids concentration was 80% of TS. The sludge samples were treated in batch in an O3 reactor schematically shown in Figure 1. As shown in Figure 1, the sludge ozonation system consists mainly of an O3 generator, an analytic system of O3 W

Figure 1

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concentration, an ozone contactor, an internal recycling system and an absorber system of exhaust. O3 gas was generated by an ozone generator (XZF-5, Unisplendour Co., Ltd) with drying air coming from the air oil-less compressor (ZWD-0.05/7, Anshan Jiapeng Compressor Co., Ltd) and was introduced into the stainless steel ozone contactor (inner diameter 10 cm and height 60 cm) through the sand core aerator. The aerator had an average pore size of 0.1 mm and the subsequent mean size of bubbles produced was 7.0 mm. The volume of the mixture of air and ozone going to the ozone contactor was measured by an anticorrosion-type wet gas ﬂow meter (LMF-1, Changchun Automotive Filter Co., Ltd), and was 0.1 m3/h (at 20 C and 1.0 atm). The ozone concentration in gas phase was determined to be 14–17 mg/L by iodometry method according to the standard procedure of by the International Ozone Association (Rakness et al. ). In order to minimize error caused by the foaming problem, an internal recycling system was designed, where the froth was broken in a stainless steel column (inner diameter 30 cm and height 60 cm) and recycled to the ozone contactor by a peristaltic pump (BT100–2J, Langer Pump Co., Ltd) at a speed of 273 mL/min. The exhaust was absorbed by potassium iodide liquor (2%) before emission to the atmosphere. The ozone dose was deﬁned as the ratio of the mass of ozone introduced into the ozonation reactor to the mass of the sludge before ozonation. W

Sludge treatment by O3/H2O2 Sludge treatment by O3/H2O2 was operated under the same condition as O3 treatment including the same quality of sludge samples and the same ozone. Different doses of

Schematic diagram of sludge oxidation process. (1) Air oil-less compressor, (2) ozone generator, (3) ozone contactor, (4) buffer bottle, (5) peristaltic pump, (6) analytic system of outlet concentration, (7) analytic system of inlet concentration, (8) absorber of exhaust.

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H2O2 were put into the reactor by peristaltic pump together with the introduction of O3. The H2O2 dose was defined as the ratio of the mass of hydrogen peroxide introduced into the reactor to the mass of the sludge before oxidation. The sludge samples without any treatment were chosen as control. Analytical methods With the disintegration of sludge, the chemical characteristics of sludge were changed greatly especially for the amount of C, N and P compound. Soluble chemical oxygen demand (SCOD) and soluble carbohydrate in sludge liquid phase were chosen to indicate the changes of carbon compounds of sludge, while total nitrogen (TN), ammonia (NHþ 4 -N), nitrite (NO2 -N) and nitrate (NO3 -N) in sludge liquid phase were selected as the indicator of nitrogen compound in sludge. In addition, the total phosphorus (TP) and reactive phosphorus (PO3 4 ) in the supernatant of sludge were also measured to reflect the changes of phosphorus compounds in sludge. The SCOD in supernatant was measured by means of the dichromate method (SEPA ). Soluble carbohydrate was measured by phenol–sulfuric acid method (Townshend et al. ) and the standard curve was made with glucose. 3 TN, NHþ were deter4 -N, NO2 -N, NO3 -N, TP and PO4 mined in accordance with standard methods (SEPA ). In this part, the samples of supernatants used for the measurement of SCOD, soluble carbohydrate, TN, NHþ 4 3 N, NO 2 -N, NO3 -N, TP and PO4 were obtained by centrifuging at 4,500 rpm for 20 min and filtration with 0.45 μm filters.

RESULTS AND DISCUSSION Carbon compound During the dissolution of sludge, carbonaceous organic substances would be released to the sludge supernatant in the form of SCOD, which includes some easily oxidized substance such as carbohydrate. So, in this work, SCOD and carbohydrate were investigated before and after sludge treatment and the results are shown in Figure 2. It can be seen that in both O3 and O3/H2O2 systems, the content of SCOD in the supernatant of sludge increased along with the addition of oxidant. However, for the same dose of O3, the content of SCOD released in the O3/H2O2 system was 1.27–1.46 times as much as that in the O3 system, showing

Figure 2

Comparison of SCOD and carbohydrate release by ozone alone and by O3/H2O2.

the effectiveness of sludge disruption by the combined O3/H2O2 system was greater than that of O3 alone. But the changes of carbohydrate in sludge supernatant were very different from those of SCOD, as there was a peak in both O3 system and O3/H2O2 systems. Especially, when O3 dose was 137.2 mg/gTS, the released carbohydrate by O3/H2O2 treatment was 1.47 times as much as that released in O3 system, which was 72% of the maximum of O3 system at 203.3 mg/gTS dose. With the increasing dose of oxidant, the amount of carbohydrate in O3/H2O2 system decreased in a faster manner than in O3 system alone. This phenomenon may be explained in two aspects. On the one hand, the dissolution of sludge could be improved greatly by adding an appropriate amount of H2O2 to the O3 system, attributable to the generation of more oxidative hydroxyl radical. Therefore, more sludge could be decomposed efﬁciently in the O3/H2O2 system, which led to a larger amount of SCOD and carbohydrate appearing in the supernatant of sludge than that in O3 system alone. However on the other hand, some easily oxidized substance such as carbohydrate in the supernatant of sludge could also be oxidized by hydroxyl radical due to its high oxidative property on all organics. So the amount of SCOD and carbohydrate in sludge supernatant was decided by the compromise of these two functions. According to the theory of sludge reduction, the minimization of sludge quantity was mainly attributed to two aspects: dissolving the sludge directly by the oxidative chemical agents and the metabolism of biodegradable substances coming from the soluble sludge by the microorganisms of the subsequent biological system (Chu et al. b). Obviously, if the second pattern could be enhanced, the cost of sludge reduction would decrease effectively. Carbohydrate was just this kind of substance that can be metabolized by microorganisms easily. So, it is very

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important to increase and maintain the content of carbohydrate in sludge supernatant during the disintegration of sludge, and a certain amount of H2O2 would facilitate this.

Nitrogenous substance As a key element to living organisms, protein is abundant in the cytoderm of Gram-negative bacteria as well as in the cell membrane and cytoplasm of bacteria, and accounts for nearly 50% of the dry material of cells (Tania et al. ). So, the content and the conversion of protein caused by sludge dissolution would have a direct effect on the performance of the subsequent biological system. Therefore, this study investigated the TN during the disintegration of sludge which was treated by O3 alone and by O3/H2O2 system. The results in Figure 3 showed that TN in the supernatant was increased markedly after sludge dissolution. Speciﬁcally, when O3 dose was 137.2 mg/gTS, the TN concentration in the supernatant rose rapidly to 59 mg/L, while in O3/H2O2 system, it reached 63 mg/L, which was 1.15 times as much as that in O3 system and accounted for 70.3% of that in O3 system when O3 dose was 203.3 mg/gTS. In wastewater the TN is considered as the sum of organic nitrogen, ammonia, nitrate and nitrite. So, the distribution and the conversion of different forms of nitrogen compounds during the whole treatment process were also investigated. Figure 4 shows the distribution of nitrogen in the supernatant of sludge which was treated by O3 alone with 290.5 mg/gTS dose and by O3/H2O2 with 290.5 mg/ gTS of O3 and 436 mg/gTS of H2O2. It was clear that organic nitrogen was the main form in the supernatant of sludge, which amounts to 82.82% of TN in O3 system and 80.6% of TN in O3/H2O2 system, while for inorganic nitrogen, ammonia was the main form and nitrate follows. Speciﬁcally, ammonia accounted for 14.9% of TN in O3

Figure 3

Comparison of TN release by ozone alone and by O3/H2O2.

Figure 4

Distribution of nitrogen in supernatant of sludge treated by O3 and by O3/H2O2.

system and 17.3% of TN in O3/H2O2 system, ranking at the second position. These results show that the reductive form of nitrogen is the overwhelmingly main form of TN, implying that the sludge cells as well as the released carbohydrate are more reductive and preferentially oxidized. Phosphoric substance Phosphorus is another important element in the composition of the organism and also a vital constituent of the cell membrane (Saktaywin et al. ). With the disintegration of sludge, the phosphorus of the solid phase was dissolved into the liquid phase of supernatant. The changes of phosphorus not only reﬂected the rupture of cells but also had a fundamental effect on the subsequent biological treatment. Figure 5 describes the changes of TP and orthophosphate (PO3 4 ) in the supernatant of sludge before and after treatment by O3 alone and by O3/H2O2. In raw sludge, phosphorus mainly existed in the solid phase and the content in supernatant was very small. In both the O3 system and O3/H2O2 systems, TP and PO3 4 in sludge supernatant increased rapidly as the reaction continued. When ozone dose was 203.3 mg/gTS, the value of TP and PO3 in supernatant of O3 system was 14.39 and 4 7.89 mg/L respectively, while in O3/H2O2 system, they were 17.29 and 8.29 mg/L respectively. After that, they

Figure 5

Comparison of TP and PO3 4 release by ozone alone and by O3/H2O2.

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increased at a lower speed in the two systems. During the whole process of reaction, the content of TP and PO3 4 in O3/H2O2 system was always higher than that in O3 system. According to the study of Dytczak, the phosphatide that constitutes the cell membrane was oxidized ﬁrst and thus led to the breakup of cytoderm (Dytczak et al. ). As a result, a large amount of polyphosphates was released from the body of phosphorus-accumulating organisms and the concentration of TP and PO3 in sludge supernatant 4 increased continuously. Consequently, the results of carbonaceous, nitrogenous and phosphoric compounds are consistent, as they all released more rapidly by the combined O3/H2O2 effect than O3 treatment alone, showing the advantage of the O3/H2O2 sludge dissolution system. This agrees well with the literature that O3 can only induce hydroxyl radicals in alkaline solutions, while adding H2O2 into O3 system would enhance hydroxyl radical formation in relatively neutral conditions, such as pH 7.03 of the sludge. Although some carbohydrate can be mineralized by the oxidants, most are still remaining and can be utilized as a carbon source of denitriﬁcation by denitrifying bacteria, thus reducing the cost of sludge treatment. So, it would be good for the subsequent biological treatment to raise the quantity of biodegradable substance represented by carbohydrate and reduce the content of the nitrogenous and phosphoric substances as far as possible. From this point of view, sludge treated by O3/H2O2 with lower ozone dose would be more appropriate for cryptic growth and to reduce the cost of sludge reduction than ozone alone.

CONCLUSIONS The effects of sludge decomposition caused by O3 alone and by O3/H2O2 were compared based on the measurement of carbonaceous organic materials, nitrogenous compounds and phosphoric substances in sludge supernatant. At the same O3 dose, the concentrations of SCOD, TN and TP in the O3/H2O2 system were higher than those in the O3 system. And thus it can be concluded that the addition of H2O2 would help to promote the capability of sludge disintegration by O3. The organic nitrogen and organic phosphorus was the main constituent of TN and TP in the supernatant of sludge in both systems. But for the substances that were easy to be oxidized such as carbohydrate, a peak would appear in both the O3 system and the O3/H2O2 system, which is the compromise of the carbohydrate release from dissolved cells and mineralization by the oxidants. All

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the SCOD, TN and TP would be further reduced by the following biodegradation system, and thus the H2O2-enhanced O3 treatment system has more potential for sludge reduction.

ACKNOWLEDGEMENTS This study was supported by the National Natural Science Foundation of China (No. 51208226) and by the project of 12th Five Year Plan in the ﬁeld of science and technology (Jilin Department of Education, No. 234).

REFERENCES Chu, L. B., Wang, J. L., Wang, B., Xing, X. H., Yan, S. T., Sun, X. L. & Jurcik, B. J. M. a Changes in biomass activity and characteristics of activated sludge exposed to low ozone dose. Chemosphere 77 (2), 269–272. Chu, L. B., Yan, S., Xing, X. H., Sun, X. L. & Jurcik, B. b Progress and perspectives of sludge ozonation as a powerful pretreatment method for minimization of excess sludge production. Water Research 43 (7), 1811–1822. Dytczak, M. A., Londry, K. & Siegrist, H.  Extracellular polymers in partly ozonated return accivated sludge: impact on ﬂocculation and dewaterability. Water Science and Technology 54 (9), 155–164. Hoigne, J. & Bader, H.  Role of hydroxyl radical reactions in ozonation processes in aqueous solutions. Water Research 10 (5), 377–386. Kiyomi, A., Terunobu, S. & Toshihiro, T.  Veriﬁcation of sludge reduction by ozonation with phosphorus recovery process at a demonstration plant. Ozone: Science & Engineering 33, 171–178. Labelle, M. A., Ramdani, A., Deleris, S., Gadbois, A., Dold, P. & Comeau, Y.  Ozonation of endogenous residue and active biomass from a synthetic activated sludge. Water Science and Technology 63 (2), 297–302. Lan, W. C., Li, Y. Y., Bi, Q. & Hu, Y. Y.  Reduction of excess sludge production in sequencing batch reactor (SBR) by lysiscryptic growth using homogenization disruption. Bioresource Technology 134, 43–50. Liu, Y.  Chemically reduced excess sludge production in the activated sludge process. Chemosphere 50, 1–7. Ludovico, S., Azize, A., Jean, C. B., Roberto, C., Pavel, J., Angelique, L., Wim, R., Guo, R. X. & Lex, V. D.  Sustainable and innovative solutions for sewage sludge management. Water 3 (2), 702–717. Ma, H. J., Zhang, S. T., Lu, X. B., Xi, B., Guo, X. L., Wang, H. & Duan, J. X.  Excess sludge reduction using pilot-scale lysis-cryptic growth system integrated ultrasonic/alkaline disintegration and hydrolysis/acidogenesis pretreatment. Bioresource Technology 116, 441–447.

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Natasa, N. T., Darja, M., Mojca, R., Lztok, A., Matja, M., Magda, C., Albin, P. & Venceslav, K.  Manganese functionalized silicate nanoparticles as a Fenton-type catalyst for water puriﬁcation by advanced oxidation processes (AOP). Advanced Functional Materials 22 (4), 820–826. Rakness, K., Gordon, R., Langlais, B., Masschelein, W., Matsumoto, N., Richard, Y., Robson, C. M. & Somiya, I.  Guideline for measurement of ozone concentration in the process gas from an ozone generator. Ozone: Science & Engineering: The Journal of the International Ozone Association 18 (3), 209–229. Saktaywin, W., Tsuno, H. & Nagare, H.  Operation of a new sewage treatment process with technologies of excess sludge reduction and phosphorus recovery. Water Science and Technology 54 (6), 424–430. Salsabil, M. R., Laurent, J., Casellas, M. & Dagot, C.  Techno-economic evaluation of thermal treatment, ozonation and sonication for the reduction of wastewater biomass volume before aerobic. Journal of Hazardous Materials 174, 323–333.

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SEPA (State Environmental Protection Administration of China)  Standard Methods for the Examination of Water and Wastewater. China Environmental Science Press, Beijing, China. Tang, Y., Yang, Y. L., Li, X. M., Yang, Q., Wang, D. B. & Zeng, G. M.  The isolation, identiﬁcation of sludge-lysing thermophilic bacteria and its utilization in solubilization for excess sludge. Environmental Technology 33 (8), 961–966. Tania, D., Liu, Y. J. & Ramesh, G.  Evaluation of simultaneous nutrient removal and sludge reduction using laboratory scale sequencing batch reactors. Chemosphere 76 (5), 697–705. Townshend, A., Worsfold, P., Haswell, S., Werner, H. & Wilson, I.  Encyclopedia of Analytical Science. Vol. 1, Academic Press, San Diego, CA, USA. Zhao, Y. X., Yin, J., Yu, H. L. & Jiang, B. J.  Observations on ozone treatment of excess sludge. Water Science and Technology 56 (9), 165–175. Zhao, Y. X., Yin, J., Yu, H. L., Jiang, B. J. & Jiang, S. K.  Sludge treatment by O3/H2O2 and carbon source recovery. Journal of Residuals Science & Technology 6 (4), 193–199.

First received 14 January 2014; accepted in revised form 2 April 2014. Available online 26 April 2014

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Pilot trial study of a compact macro-ﬁltration membrane bioreactor process for saline wastewater treatment Dao Guan, W. C. Fung, Frankie Lau, Chao Deng, Anthony Leung, Ji Dai and G. H. Chen

ABSTRACT Conventional membrane bioreactor (MBR) systems have increasingly been studied in recent decades. However, their applications have been limited due to their drawbacks such as low flux, membrane fouling, and high operating cost. In this study, a compact macro-filtration MBR (MfMBR) process was developed by using a large pore size membrane to mitigate the membrane fouling problem. A pilot trial of MfMBR process was set up and operated to treat 10 m3/day of saline wastewater within 4 h. The system was operated under an average permeate flux of 13.1 m3/ (m2·day) for 74 days. The average total suspended solids, total chemical oxygen demand, biological oxygen demand, total Kjeldahl nitrogen, and total nitrogen removal efficiencies achieved were 94.3, 83.1, 98.0, 93.1, and 63.3%, respectively, during steady-state operation. The confocal laser scanning microscopy image indicated that the backwash could effectively remove the bio-cake and dead bacteria. Thus, the results showed that the MfMBR process, which is essentially a primary wastewater treatment process, had the potential to yield the same high quality effluent standards as the secondary treatment process; thereby suggesting that it could be used as an option when the economic budget and/or land space is limited. Key words

Dao Guan Chao Deng Anthony Leung Ji Dai G. H. Chen (corresponding author) Department of Civil and Environmental Engineering, Hong Kong University of Science & Technology, Clear Water Bay, Kowloon, Hong Kong, China E-mail: ceghchen@ust.hk W. C. Fung Frankie Lau Drainage Services Department, Government of the Hong Kong Special Administration Region, 44/F, Revenue Tower, 5 Gloucester Road, Wanchai, Hong Kong, China

| marco-ﬁltration membrane bioreactor, organic and nitrogen removal, pilot trial, saline wastewater

INTRODUCTION In recent decades, membrane bioreactor (MBR) technologies have been developed and applied to water and wastewater treatment ( Judd & Judd ). Compared with the conventional activated sludge process, MBR technology has many advantages including small footprint, low excess sludge production, and good effluent quality (Melin et al. ). Additionally, due to its long sludge retention time (SRT) and complete separation between hydraulic retention time (HRT) and SRT, the MBR process can achieve high ammonium removal efficiency at a short HRT (Henze et al. ). However, its application has mainly been limited to high-strength industrial wastewater or domestic wastewater treatment for effluent reuse (Gander et al. ). Membrane fouling control and low permeate flux are the major constraints (Le-Clech et al. ), which result in relatively high energy consumption and high operating cost (Zhang et al. ; Judd ). Moreover, doi: 10.2166/wst.2014.180

membrane replacement further increases the maintenance cost (Meng et al. ). In order to ease these constraints, a high-flux macrofiltration MBR (MfMBR), with a mean pore size >5 μm, has been developed in Hong Kong (Chen & Pang ). This MfMBR process achieved effective permeate flux of 6 m3/(m2·day) under the relatively low air-to-water ratio of 15 (Bai et al. ). A pilot trial of this process for saline sewage treatment demonstrated a non-fouling operation period of up to 370 days at such high permeate flux (Bai et al. ). This paper reports on the pilot trial of an optimized MfMBR targeting conditions of higher flux and shorter HRT than presented in most literature. These conditions may be suitable for the upgrade of existing chemically enhanced primary treatment (CEPT) plants towards secondary treatment with minimized land space and sludge disposal requirements. The MfMBR process

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presented in this study integrates a sludge hydrolysis tank under oxic–anaerobic conditions that are similar to those of the oxic–settling–anoxic (OSA) process developed by Chen et al. (). Hence, this tank is named OSA tank, whose functions are to stabilize the MfMBR sludge, reduce its sludge production and generate secondary substrate (mainly volatile fatty acids, VFAs) to enhance post-denitriﬁcation. The whole process is named the membrane enhanced primary treatment (MEPT®) process because its total HRT of 4 h is comparable to that of the CEPT process.

MATERIALS AND METHODS MEPT pilot plant A MEPT pilot plant was installed and constructed at Kwai Chung industrial wastewater pumping station, which receives about 10 m3/day of wastewater (Figure 1). The process consisted of two tanks: an aerobic MfMBR tank and an OSA tank. The MfMBR tank was a cuboid tank having a membrane set and aeration diffusers; the volume of the tank was 0.833 m3. The membrane set consisted of 10 membrane modules, which provided a total effective membrane surface area of 0.184 m2. The membrane material was nylon woven fabric mesh whose mean pore size was approximately 55 μm. The OSA tank was a 0.833 m3 anoxic cuboid tank with a propeller mixer. Raw wastewater was fed into the membrane tank after passing through a 2 mm ﬁne screen, and subsequently ﬂowed through an equalization tank having a short HRT of

Figure 1

Schematic of the MEPT pilot plant.

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14 min. After 2 h of aerobic reaction in the MfMBR tank, the mixed liquor entered the OSA tank for another 2 h of anoxic–anaerobic reaction. The flow recirculation ratio between these two tanks was set at 1. The dissolved oxygen (DO) concentration in the aerobic tank was kept higher than 2 mg/L and that in the OSA tank was below 0.5 mg/L (data not shown). The effluent was continuously drawn from the MfMBR tank under an operational permeate flux of 13.1 m3/(m2·day). Daily membrane backwash with the effluent of the system was programmed to last 1 min every 24 h. Additionally, there was no chemical cleaning during the entire trial. The operation of the entire plant was controlled via an automatic programmable logic controller equipped with water level sensors. The seeding sludge was taken from the Shatin sewage treatment works of Hong Kong, which uses the conventional activated sludge process for the treatment of mainly domestic wastewater. The mixed liquor suspended solids (MLSS) concentration in both tanks was maintained at around 3.5 g/L. Throughout the whole trial period no sludge withdrawal was purposefully done. Sampling and sample analysis From March to May 2012, 24 h composite samples of the influent, effluent, and mixed liquor of the MBR and OSA tanks were collected every 2 days by using a peristaltic pump with automatic control. All samples were stored at 4 C. The 24 h composite samples for total chemical oxygen demand (TCOD) and soluble chemical oxygen demand (SCOD) were analyzed on site, while the 24 h W

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composite samples for biological oxygen demand (BOD5) total suspended solids (TSS), ammonium (NHþ 4 ), total Kjeldahl nitrogen (TKN), and total nitrogen (TN) were measured in the laboratory within 12 h of sampling; so also were the grab samples for MLSS. At the end of the trial several pieces of membrane were cut off for bio-cake analysis, and subsequently preserved at 4 C. The membrane pieces were then transported to the laboratory, and dyed and then examined by confocal laser scanning microscopy (CLSM) within 2 h. W

Analytical methods The determination of BOD5 followed the British Standard BS EN 1899-1:1998 (BSI ), according to which nitrification is inhibited by allylthiourea (ATU). The procedures for measuring MLSS, TSS, TCOD, SCOD, TKN, ammonium, nitrate (NO 3 ), nitrite (NO2 ), chloride (Cl ), sulfate 2 (SO4 ), DO concentrations, pH, and temperature followed Standard Methods for the Examination of Water and Wastewater (APHA ). Prior to the measurement of ammonium, nitrate and nitrite, samples were filtered through 0.45 μm filter membrane and then concentrations measured with a flow injection analyzer (FIA) (QuickChem FIA þ 8000Series, Lachat, USA). The TKN samples were digested without filtration, before subjecting them to the FIA analysis. The TN concentration was determined from the TKN, nitrate, and nitrite concentrations. Chloride and sulfate concentrations were detected using ion chromatography (LC-20AD SP, Shimadzu, Japan) with a 7 μm pore AllsepTM A-2 anion column (Alltech, USA). The DO concentration was measured with a DO meter (YSI 52, USA). The TCOD and SCOD concentrations were determined with a COD reactor (HACH) coupled with a direct reading spectrophotometer (DR/2000, HACH, USA). In order to investigate the fouling conditions, when a membrane module was removed from the MfMBR tank, several pieces of the membrane were cut off for bio-cake examination, including application of the LIVE/DEAD® BacLight™ Bacterial Viability and Counting Kit (L34856, Life Technologies, USA), to visualize the live and dead bacteria cells, in which green-fluorescent SYTO®9 and red-fluorescent propidium iodide were applied as dyes in sequence. After 15 min of incubation in the dark, samples were examined using CLSM (Zeiss, LSM7 DUO [710 þ LIVE], Germany). More than nine images at different locations of each sample were taken to ensure the representativeness.

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RESULTS AND DISCUSSION Influent condition The pilot plant suffered from high salinity in the influent due to the application of seawater toilet flushing in Hong Kong. The average Cl and SO2 4 concentrations during the trial period were 6,473.84 ± 1,514.65 and 133.77 ± 71.49 mg/L, respectively. The detailed characterization of the influent is shown in Table 1. The high standard deviations of nitrate and nitrite (NOx-N) and TN concentrations of 29 ± 83 and 84 ± 93 mg/L, respectively, were interpreted as being due to the presence of a fraction of industrial wastewater in the influent, which was not expected since the MfMBR was designed for domestic wastewater treatment. However, the constant volatile suspended solids to TSS ratio, stable TSS concentration, and high ammonium and TN removal efficiencies indicated that the influent had no apparent toxic effect on the sludge; this will be discussed later. TSS removal TSS concentrations in the influent and effluent and TSS removal efficiency are shown in Figure 2. Except in the 28 days of start-up operation, the effluent TSS was always below 30 mg/L despite great variation in the influent TSS from 500 to 1,000 mg/L. The overall TSS removal efficiency was 94.3%, suggesting that the MEPT® process could achieve good TSS removal. The pore size of the membrane used was larger than 50 μm, which was in the range of the activated sludge particle size distribution. Comparing with conventional micro-filtration (MF) and ultra-filtration (UF) membranes, the bio-cake formed on the membrane surface in the MfMBR system plays a key role on filtering the activated Table 1

Average influent and effluent quality as well as removal efficiencies of the MEPT® process during 46 days of steady-state operation

Parameter

Inﬂuent

Efﬂuent

Removal efﬁciency (%)

TSS (mg/L)

586 ± 187

28.8 ± 17.5

94.3 ± 3.1

SCOD (mg/L)

336 ± 193

54.1 ± 33.6

80.3 ± 10.6

TCOD (mg/L)

593 ± 288

86.0 ± 37.0

83.1 ± 8.4

BOD5 (mg/L)

318 ± 177

4.6 ± 2.9

98.2 ± 1.4

NH3-N (mg/L)

31.6 ± 13.7

1.38 ± 1.49

95.3 ± 5.3

NOx-N (mg/L)

29 ± 83

13.9 ± 9.8

TKN (mg/L) TN (mg/L)

–

44.6 ± 18.8

2.69 ± 1.62

93.1 ± 4.6

84 ± 93

20.2 ± 9.0

63.3 ± 23.1

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Figure 2

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Pilot trial study of MfMBR for saline wastewater treatment

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3–4 mg/L. The good removal of TCOD was attributed to stable bio-cakes having formed, thus providing good solids separation. The average removal efficiencies of TCOD and BOD5 during the steady-state period were 83.1 and 98.0%, respectively. The results suggested that the process could remove almost all biodegradable COD in wastewater. Therefore, the relatively low TCOD removal efficiency and relatively high effluent TCOD concentrations were attributed to the high portion of slowly biodegradable and nonbiodegradable carbon in the raw wastewater from an industrial area, which could not be removed during the short HRT of 4 h. TSS removal performance in MEPT pilot trial.

Nitrogen removal sludge – an observation also made by Fan & Huang () who indicated that during the operation of large pore size membrane, a formation stage would be necessary for the cake layer to form on the supporting material. The high effluent TSS in the start-up period was due to incomplete formation of the bio-cake on the membrane surface. Therefore, a relatively long period was necessary for the build-up of the bio-cake layer. This indicated that a start-up period might be a requirement for the MfMBR process although its duration could be reduced depending on operational strategy such as the provision of sufficient seeding sludge. Organic carbon removal Figure 3 shows the TCOD and BOD5 concentrations in the influent and effluent and respective removal efficiencies. Despite great TCOD fluctuation in the first 28 days, the effluent TCOD became stable at around 86 mg/L after this period. Similar to TCOD, the effluent BOD5 was stable at

Figure 3

(a) BOD5 and (b) TCOD removal performance in MEPT pilot trial.

Figure 4 shows the TKN and TN removal of the MEPT pilot trial. During the start-up period, the drop in the effluent TKN, which was as a result of improved TKN removal efficiency, was expected as more nitrifiers accumulated in the system owing to well developed bio-cakes. The TN removal efficiency during the steady-state period was 63.3% even when the influent nitrate was above 200 mg/L, which is much higher than typical municipal wastewater. These results indicated that the OSA tank in the MEPT process satisfactorily removed TN. The low BOD5 level in the effluent of the MfMBR tank indicated that readily biodegradable organics, fed through the influent, were probably consumed in the aerobic tank (see Figure 3(a)). Consequently, the carbon source available for denitrification in the OSA tank became insufficient unless extra carbon was provided through the hydrolysis of sludge. Previous studies (Saby et al. ; Yuan et al. ) reported that sludge decay could take place in the OSA tank, thus generating VFAs, which became the carbon source for denitrification.

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Figure 4

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Pilot trial study of MfMBR for saline wastewater treatment

In general, the MEPT® process performed well during the steady-state operation. A summary of the quality of influent and effluent as well as the removal efficiencies is given in Table 1. The rather large standard deviation indicated that the quality of the influent raw wastewater varied significantly, especially its NOx-N concentration. However, the effluent quality and removal efficiency were both relatively stable during the 46 days of steady-state operation. In summary, the overall performance of the MEPT® process was satisfactory in terms of membrane filtration, biological conversion, organics, and nitrogen removal. The effluent quality was consistently good and could fulfill the effluent discharge standards in Hong Kong (EPD ). The removal efficiencies in terms of TSS, TCOD, BOD5, ammonia, TKN, and TN were high. Although the effluent SCOD concentration was not low, the removal of biodegradable carbon in the MEPT® process was high as

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(a) TKN and (b) TN removal performance in MEPT pilot trial.

General performance of MEPT® process

Figure 5

Water Science & Technology

indicated by the high BOD5 removal. This showed that the bulk of the effluent SCOD might be in the form of unbiodegradable organics. The average MLSS concentrations in the MfMBR and OSA tank were 3,214 and 3,188 mg/L, respectively. The calculated SRT of the MfMBR tank was 18 days. The observed overall sludge yield of the plant (including the solids loss via the effluent) was approximately 0.1 gMLSS/gCOD, which is much lower than that of the conventional anoxic/oxic process (i.e. around 0.35 gMLSS/gCOD when the SRT is 20 days) (Muller et al. ). Membrane condition During the 74 days of operation, the MfMBR system maintained a high permeate flux with an average value of 13.1 m3/(m2·day), while the backwash frequency was maintained at 1 min every 24 h. The operational flux in this study was significantly higher than other studies

CLSM image analysis on the bio-cake formed on the membrane before (a) and after (b) backwash at the end of the trial. Grey (green in online version) indicates live bacteria. Dead bacteria appear as black (red in online version) color. The full color version of this ﬁgure is available online at http://www.iwaponline.com/wst/toc.htm.

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whereas the membrane pore size was in the same range (0.4–5 m3/(m2·day)) (Ersahin et al. ). Nonetheless, membrane fouling was under control as indicated by the low operational transmembrane pressure (TMP) value (<0.2 bars). At the end of the trial, live/dead double-staining combined with the CLSM image analysis was conducted to examine the characteristics of the bio-cake on the membrane surface. Figure 5(a) shows that the bio-cake developed well on the membrane surface, thus suggesting that a well-formed bio-cake could act as a filter to achieve the good effluent quality obtained in this study. Moreover, dead bacteria accumulation on the membrane surface was negligible, as indicated by the few black (red in online version) dots. The accumulation of dead bacteria has the potential to increase TMP as well as membrane fouling (Hwang et al. ). However, the fact that no apparent membrane fouling occurred in this study suggested that no or only few dead bacteria accumulated on the membrane. After membrane backwash, the bio-cake had significantly been removed, as shown in Figure 5(b), which suggested that backwash was still effective at the end of the trial, hence further substantiating the non-fouling condition of the membranes.

CONCLUSIONS In this study, the performance of the MEPT® process was closely monitored and investigated under conditions of high flux and short SRT. The major findings are as follows. (1) The pilot plant can remove TSS effectively and generate effluent with high quality comparable to that of the secondary wastewater treatment process. Compared with traditional MF or UF membrane, the MfMBR can provide an alternative solution when the effluent discharge standard (EPD ) is not rigorous and the budget is limited. (2) Under a high flux (13.09 m3/(m2·day)), low backwash frequency (1 min every 24 h) and no chemical cleaning operation, the membranes operated continuously for 74 days without membrane fouling. (3) The MEPT® process can achieve remarkable TKN removal, high COD and BOD removal, and a good TN removal efficiency. (4) The OSA tank produced enough carbon source, in the form of VFAs, from sludge hydrolysis under wellcontrolled conditions to enhance TN removal.

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ACKNOWLEDGEMENT This study was supported by the Drainage Services Department of the Government of the Hong Kong Special Administration Region (Project no. DEMP10/10).

REFERENCES APHA  Standard Methods for the Examination of Water and Wastewater, 21st edn. American Public Health Association/ American Water Works Association/Water Environment Federation, Washington, DC, USA. Bai, P., Chen, X., Yin, H., Wang, J. & Chen, G. H.  A real trial of an innovative membrane bioreactor for saline sewage treatment. Desalination and Water Treatment 18 (1–3), 297–301. Bai, P., Wang, J. & Chen, G. H.  A real trial of a long-term nonfouling membrane bioreactor for saline sewage treatment. Water Science and Technology 63 (7), 1519–1523. BSI  BS EN 1899-1 Water Quality–Determination of Biochemical Oxygen Demand after n Days (BODn) – Part 1: Dilution and Seeding Method with Allylthiourea Addition. British Standards Institution, London. Chen, G. H. & Pang, S. K.  Development of a low-cost dynamic filter immersed in activated sludge system. Water Science and Technology: Water Supply 6 (6), 111–117. Chen, G. H., An, K.-J., Saby, S., Brois, E. & Djafer, M.  Possible cause of excess sludge reduction in an oxic-settlinganaerobic activated sludge process (OSA process). Water Research 37 (16), 3855–3866. EPD  Technical Memorandum Standards for Effluents Discharged into Drainage and Sewerage Systems, Inland and Coastal Waters. EPD, Hong Kong, http://www.legislation. gov.hk/blis_pdf.nsf/6799165D2FEE3FA94825755E0033E 532/569E03D57CCBAE69482575EE006FF774/$FILE/ CAP_358AK_e_b5.pdf (accessed 10 March 2014). Ersahin, M. E., Ozgun, H., Dereli, R. K., Ozturk, I., Roest, K. & van Lier, J. B.  A review on dynamic membrane filtration: materials, applications and future perspectives. Bioresource Technology 122, 196–206. Fan, B. & Huang, X.  Characteristics of a self-forming dynamic membrane coupled with a bioreactor for municipal wastewater treatment. Environmental Science and Technology 36 (23), 5245–5251. Gander, M., Jefferson, B. & Judd, S.  Aerobic MBRs for domestic wastewater treatment: a review with cost considerations. Separation and Purification Technology 18 (2), 119–130. Henze, M., Van Loosdrecht, M. C. M., Ekama, G. A. & Brdjanovic, D.  Biological Wastewater Treatment: Principles, Modeling, and Design. International Water Association, London, UK. Hwang, B.-K., Lee, W.-N., Yeon, K.-M., Park, P.-K., Lee, C.-H., Chang, I.-S., Drews, A. & Kraume, M.  Correlating TMP increases with microbial characteristics in the bio-cake on

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the membrane surface in a membrane bioreactor. Environmental Science and Technology 42 (11), 3963–3968. Judd, S.  The status of membrane bioreactor technology. Trends in Biotechnology 26 (2), 109–116. Judd, S. & Judd, C.  The MBR Book – Principles and Applications of Membrane Bioreactors for Water and Wastewater Treatment. Elsevier, Oxford, UK. Le-Clech, P., Chen, V. & Fane, T. A. G.  Fouling in membrane bioreactors used in wastewater treatment. Journal of Membrane Science 284 (1–2), 17–53. Melin, T., Jefferson, B., Bixio, D., Thoeye, C., De Wilde, W., De Koning, J., van der Graaf, J. & Wintgens, T.  Membrane bioreactor technology for wastewater treatment and reuse. Desalination 187 (1–3), 271–282. Meng, F., Chae, S.-R., Drews, A., Kraume, M., Shin, H.-S. & Yang, F.  Recent advances in membrane bioreactors (MBRs):

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membrane fouling and membrane material. Water Research 43 (6), 1489–1512. Muller, A., Wentzel, M. C., Loewenthal, R. E. & Ekama, G. A.  Heterotroph anoxic yield in anoxic aerobic activated sludge systems treating municipal wastewater. Water Research 37 (10), 2435–2441. Saby, S., Djafer, M. & Chen, G. H.  Effect of low ORP in anoxic sludge zone on excess sludge production in oxicsettling-anoxic activated sludge process. Water Research 37 (1), 11–20. Yuan, Q., Sparling, R. & Oleszkiewicz, J. A.  VFA generation from waste activated sludge: effect of temperature and mixing. Chemosphere 82 (4), 603–607. Zhang, J., Chua, H. C., Zhou, J. & Fane, A. G.  Factors affecting the membrane performance in submerged membrane bioreactors. Journal of Membrane Science 284 (1–2), 54–66.

First received 16 January 2014; accepted in revised form 31 March 2014. Available online 26 April 2014

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Linking the management of urban watersheds with the impacts on the receiving water bodies: the use of ﬂow duration curves Guido Petrucci, Fabrice Rodriguez, José-Frédéric Deroubaix and Bruno Tassin

ABSTRACT There is growing evidence that changes in the current hydrological behaviour of urbanising catchments are a major source of impacts on the downstream water bodies. However, current flow-rates are rarely considered in studies on urban stormwater management, usually focused on extreme flow-rates. We argue that taking into account receiving water bodies is possible with relatively small modifications in current practices of urban stormwater modelling, through the use of Flow duration curves (FDCs). In this paper, we discuss advantages and requirements of the use of FDCs. Then, we present an example of application comparing source control regulations over an urbanised catchment (178 ha) in Nantes, France. Key words

| ﬂow duration curves, model calibration, source control, stormwater management, urban hydrology

Guido Petrucci (corresponding author) José-Frédéric Deroubaix Bruno Tassin Université Paris-Est, LEESU, UMR MA102, UPEC, ENPC, UPEMLV, AgroParisTech, 6-8 avenue Blaise Pascal, 77455 Marne-la-Vallée cedex 2, France E-mail: guido.petrucci@vub.ac.be Guido Petrucci Vrije Universiteit Brussel (VUB), Earth System Sciences (ESSc), Pleinlaan 2, 1050 Brussels, Belgium Fabrice Rodriguez LUNAM Université, IFSTTAR, GER, Bouguenais, France

INTRODUCTION A growing concern for urban stormwater management is related to its impact on the receiving water bodies. Impacts can range from acute or chronic pollution to streams’ erosion or lakes’ eutrophication or contamination, with severe consequences on both the possible uses of water bodies and their environmental value (Walsh et al. b; Jung et al. ). The necessity to account for these impacts when planning or managing urban stormwater systems is widely recognised by scientists, technicians and decision-makers. However, this recognition is much more theoretical than practical: in current practices of urban hydrological studies, receiving water bodies are often poorly considered. This point could be illustrated using the example of stormwater source control. Source control (also called low impact development) mainly consists of the implementation of small-scale stormwater facilities (best management practices, or BMPs) over an urban catchment, upstream of the drainage system. One principle of source control is to mimic the pre-development behaviour of the catchment, doi: 10.2166/wst.2014.206

thus minimising the impacts of urbanisation on the downstream environment (Booth & Jackson ). This strategy is mainly implemented through source control regulations, demanding to include BMPs in any new urban development project (Balascio & Lucas ). A study on source control regulations (Petrucci ; Petrucci et al. ), showed that the preservation of downstream water bodies is a commonly stated objective of local authorities adopting these regulations. Still, when technical studies are done to define regulations, in most cases the only indicators considered are the peak flow-rates for extreme rainfall events (e.g. return periods T 10 years). This attention on extreme flow-rates indicates the persistence of a conventional approach, in which the only objective is to minimise sewer overflows and flooding, regardless of downstream water bodies. In this paper are analysed the current modelling practices in urban stormwater management and their capability to describe the effects of urbanisation on the downstream environment. On this basis, a viable approach

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is proposed to integrate downstream environment considerations into current modelling practices, using ﬂow duration curves (FDCs). This proposal is then discussed and illustrated by an example about source control regulations.

Current trends in modelling urban stormwater systems There are three main modelling approaches in current urban hydrology studies: flow-rate modelling, quality modelling and integrated modelling. Flow-rate modelling is the most classical approach: it consists of modelling the urban watershed and the drainage system in order to simulate water flows. This approach has a long tradition, and nowadays is mainly represented by distributed, detailed models (e.g. SWMM 5, MIKE URBAN). Many local authorities have implemented this kind of model for the urban areas they administer. They are used to determine interventions on the sewer system (Gironàs et al. ), or for more sophisticated purposes like real-time control (Schütze et al. ). The main limit in the current use of these models for integrating downstream water bodies is that these traditional models, despite their large possibilities, are often used in a narrow way. Often, they are calibrated on a few rain events and applied just to simulate single events or short time series supposed to benchmark the functioning of the sewer system. In general, nothing inside these models constrains their application to important rain events: their structure is able to simulate low flows as well as high and current behaviours of the catchment as well as extreme. Most current models are able to simulate long time series instead of single rain events (Elliott & Trowsdale ). The second approach is water quality modelling. It allows simulation of the flow of pollutants from the urban area. In principle, this type of modelling is extremely important to assess the impact on downstream water bodies, and its diffusion is an encouraging trend. Also because of legal constraints on stormwater quality, most flow-rate models now include water-quality simulation capabilities. However, because of a frequent lack of data, measurement uncertainties, questions on relevant physical and chemical processes, and, in general, minor experience of this type of modelling compared to flow-rate modelling, water quality models have a smaller predictive reliability than flow-rate models (Bertrand-Krajewski ). Uncertainties are amplified by a delicate modelling process, requiring significant efforts and know-how.

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The third approach is ‘integrated modelling’. This term defines the coupling of several models, each one representing a part of a complex system. Several modelling attempts belonging to this approach focus on the coupling between sewer systems and receiving water bodies. For example, Silva et al. () modelled algal blooms in an urban lake as a consequence of changes in the upstream urban drainage. This type of modelling is still the object of single case-specific research, and cannot be considered as a common practice for stormwater management purposes. Among the three approaches, the most current and solid is flow-rate modelling. However, this approach is extremely focused on sewer overflows (peak flow-rates) and does not take into account the effects downstream or those which occur due to smaller but more frequent flow disturbance. The development of other models to a level of ease of use and reliability suitable for widespread use by local authorities will probably take years. To increase the sustainability of urban water systems it is necessary to find practical and viable solutions to improve the accounting for impacts of decision-making on receiving water bodies (Ashley & Hopkinson ). For this reason, in this paper, we propose appropriate tools within the existing flow-rate modelling approach, and demonstrate their potential for application to downstream effects of urban stormwater management. Impact of urbanisation on downstream water bodies The most evident effect of urbanisation on catchment hydrology, recognised for several decades, is an increase in peak flow-rates. This phenomenon was particularly noticed for large rain events, when floods or rapid stream erosion were observed. The efforts in stormwater management were thus oriented, for a long time, to control this kind of high-flow event. Typical measures in the USA (Balascio & Lucas ) or in the UK (Faulkner ) address events with return periods higher than 1 or 2 years. In France, most stormwater systems are sized for a return period of 10 years (Petrucci et al. ). This longlasting practice of focusing only on relatively uncommon events persists in a large majority of urban hydrological studies. High-flow events surely have an impact on the receiving water bodies. However, since the 1990s a consistent literature (summarised by Walsh et al. b) started to reconsider their relevance, increasing the importance attributed to current events. In terms of channel erosion it has been shown that erosive flow-rates are small

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enough to be exceeded even by frequent rain events (Hunt & Tillinghast ). These are likely to have a stronger effect, on the long term, than large uncommon events (Booth & Jackson ). The same analysis on the increased frequency of exceedance of a ‘disturbance threshold’ because of small rain events is suggested for the biological conditions of streams (Booth et al. ; Roy et al. ). In terms of water quality, two effects should be mentioned: the first is the production of runoff, not occurring in natural catchments, for rain events of a few millimetres (Walsh et al. a). This frequent runoff, even if hydraulically negligible, may constitute a relevant intake of pollutants (and heat). The second effect is the increase in combined sewer overflows (CSOs). In conclusion, there is a growing interest for low flowrates. The behaviour of an urban catchment during extreme rain events is significant in terms of urban flooding, thus representing a point of view focused on what happens inside the urban area. If we are interested in what happens downstream, we should look at the current behaviour of the urban catchment.

CONSTRUCTION AND USE OF FDCS Consequent to the growing interest for current catchments behaviour, some authors suggested hydrologic metrics linked to downstream effects. Booth et al. () suggested using the annual fraction of days in which the daily mean discharge exceeds the annual mean discharge (TQmean). This indicator can distinguish catchments having a ‘flashy’ response from others with gradually varying flow regimes. However, the information given by such specific indicators is difficult to generalise to other geographic, climatic and urban conditions from that where the indicators were developed. That is why several researchers (Fennessey et al. ; Rohrer et al. ) used a more general way to characterise catchments’ current behaviours, using FDCs. A FDC represents the fraction of time during which a given level of flow-rate is matched or exceeded. This kind of representation has been largely employed in water management, because of its capacity to represent a huge quantity of hydrological information on a single view (Vogel & Fennessey ). Starting from a flow-rate time series, the simplest way of constructing a FDC is to order data by decreasing flow-rates, and to plot the result in frequency–flow-rate axes. Formally, starting from a series of n ordered flow-rates q(i), where i ¼ 1, … , n, the value of

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the FDC Qp, for the frequency p, is: Qp ¼ q(i) Qp ¼ q(iþ1)

if i ¼ [(n þ 1)p] if i < [(n þ 1)p]

where [(n þ 1)p] is the integer part of (n þ 1)p. This formalism, as well as more sophisticated and statistically robust methods to calculate the FDC, can be found in Vogel & Fennessey (). The work of these authors addresses also the strong dependence of FDCs on the period used for calculation, a potential weakness of this instrument. In particular, the extremes of the curve are highly dependent on the extreme values of the time series, and they are poorly reliable. The solution suggested by Vogel & Fennessey () is, for time series spanning several years, to calculate several annual FDCs instead of a unique FDC for the whole recording period. Starting from annual FDCs, they calculate a median annual FDC and its conﬁdence intervals. This method generates both more reliable FDCs and an estimation of the inter-annual variability of the hydrological regime. Another point, useful for the following discussion, is that, starting from a time series recorded at a given timestep (e.g. 5 min), it is straightforward to construct series with longer time-steps (10 min, 1 hour, 1 day), by averaging subsequent records. Thus, starting from a time series, it is possible to obtain FDCs with several time-steps. Starting from a time series, it is also possible to construct curves other than FDCs but with similar construction procedures and meaning (Fennessey et al. ).

Use and interpretation of FDCs in urban hydrology As mentioned in the introduction, most ﬂow-rate models can perform continuous simulations over long periods of time. Thus, for any scenario of stormwater management that can be described by these models, it is possible to obtain long-term simulated time series and, consequently, the corresponding FDC. Scenarios can include new infrastructures (e.g. reservoirs), new operation rules (e.g. activation rules for pumps), new regulations (e.g. source control) and new urban developments. FDCs can then be used to compare alternative scenarios. This can be done directly on the FDCs themselves (Fennessey et al. ), providing in a single view a comparison of the hydrologic regime under each scenario.

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When the comparison involves many alternatives, and/or some case-specific objective is set, it is possible to define indicators on the FDC. To facilitate comparisons, one can use generic indicators (Braud et al. ) like the flow-rates exceeded for given fractions of the year or the intersections between FDCs and the axes. Figure 1 represents a sample FDC of an urban catchment, with its confidence intervals. In the figure Q0.1%, Q1% and Q10% are indicators corresponding to the values of the FDC for frequencies of exceedance of, respectively, 0.1, 1 and 10%, while f0 is the intersection between the FDC and the x axis. The most interesting application of FDC indicators, however, lies in the possibility to define parameters linked to receiving water status on a specific case-study. The indicator TQmean proposed by Booth et al. () as explanatory of the downstream biological status can be computed from FDCs. If some ‘disturbance thresholds’ can be estimated, their frequency of exceedance can be compared for different scenarios. In the same way, the frequency of CSOs can be estimated if the overflow threshold is known. Also f0, defined above, describing the annual frequency of runoff, can be a significant indicator of water quality and biological disturbance (Walsh et al. a, b). In general, which urban impacts are relevant in each specific case

Figure 1

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highly depends on local characteristics and on the relative dimension of the urban area to the receiving water body, demanding the deﬁning of case-by-case indicators. As anticipated, it is possible to adapt the FDC time-step to the time-scales of the system studied. Models of urban stormwater systems often use short time-steps (5–10 min) to ﬁnely describe the hydraulic system. A similar time-step can be appropriate for computing CSOs or other hydraulicrelated processes, but a coarser time-step (hourly, daily) can be more adapted to describe impacts on systems having longer response-time.

Required changes in modelling practices If, in general, each scenario that can be tested in terms of peak flow-rate can be tested in terms of FDC, three main requirements have to be satisfied. The first is the availability of local rainfall data over long periods: peak flow-rate analyses could be based on few real or synthetic rainfall events, while for long-term simulations, representative inputs are necessary. However, this requirement can be easily fulfilled thanks to the rain gauge networks that have been in place for many years in most urban areas.

Examples of indicators deﬁned on an annual median FDC. Grey lines represent 90% conﬁdence intervals (15 years of simulation at 5-minute time-step). Source: Petrucci (2012).

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The second requirement is the calculation capability necessary to run long-term simulations. Still, with the exception of very large and complex urban catchments, requiring specific solutions, present calculation resources can in most cases run hydrologic simulations in reasonable times. The third and more problematic requirement is that the model should be reliable when simulating a catchment’s current behaviours. Today this is not always the case: even if flow-rate models are able to simulate the whole hydrologic regime, in practice they are often calibrated on a few high flow-rate events. This procedure focused on peaks does not guarantee the reliability of the model for low flow-rate events and for the parts of the hydrologic response other than peaks (ascending and descending limbs of the hydrograph). The reliability of the model in these terms is a primary condition for the reliability of the FDCs obtained. The solution to this reliability issue demands the probably most important evolution in current modelling practices. It requires calibrating the model on low flows as well as on high flows, and considering the whole hydrograph shape and not just the peak position and amplitude. Further, event-based calibration is not sufficient: time series of input (rainfall) and output (flow-rate) are necessary. Although rain data are easily available, flow-rate time series are less common. Calibration of flow-rate models is also complex because typical models can involve hundreds of parameters, often difficult to estimate a priori. To calibrate just on some peaks, manual calibration of a few parameters can often be satisfactory, but to fit the model for the whole catchment’s response, a better procedure is necessary (Gupta et al. ). Technical analyses of urban stormwater management should start to include automatic calibration methods, quite diffused in hydrologic research but scarcely applied in practice. Another potential difficulty in the use of FDCs can be mentioned. In most situations, the reference case for scenario comparison can be the actual state of the system. However, when the objective is to restore the ‘natural’ behaviour of a catchment, a pre-development FDC is necessary. Defining this curve can be difficult in the absence of suitable data or of a nearby reference catchment.

EXAMPLE OF APPLICATION IN URBAN HYDROLOGY Methodology We propose the application of FDC analysis to the comparison of different source control regulation scenarios. This example

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(detailed in Petrucci ) is intended here to show the better insight given by FDCs on the catchments’ behaviour. We considered the ‘Gohards’ urban catchment, in Nantes, France (Figure S1, supplementary material, available online at http://www.iwaponline.com/wst/070/206. pdf), already studied and described by Rodriguez et al. (). This catchment (178 ha) is covered by a mixed land-use (residential, commercial, industrial and agricultural), giving an impervious cover of 0.38. The sewer system is separated (14.8 km); its average slope is 0.79%. Available data include 5 years of flow-rate measurements at the outfall of the sewer system (1999–2003), and 10 years of rainfall measurements at two rain gauges inside the catchment (1999–2008). Both time series have a 5minute time-step. This catchment was chosen for this study because of the long available observation period. The model used is SWMM5, with typical methods for subcatchments’ delineation and parameters’ estimation (Gironàs et al. ). The only particular feature of the modelling setup is that we distinguished, in each subcatchment, roof, road and green areas, in order to prepare the simulation of source control regulation scenarios (Petrucci et al. ). After setup, the model was calibrated and validated. As for the model setup, all choices in the calibration/validation procedure were a compromise between accuracy and simplicity: the purpose was to reach our aim using the most current options in hydrological literature, in order to minimise the changes from current modelling practices. Calibration parameters were nine global parameters (roughness of pipes and surfaces, infiltration, initial losses, etc.) and one shape parameter (length of the overland flow path) for each subcatchment, for a total of 101 parameters. To cope with so many parameters, calibration was performed using a genetic algorithm, an optimisation algorithm largely applied in hydrology (Savic & Khu ). Optimisation consisted of maximising the commonly used Nash criterion to keep the example as simple as possible; multi-objective optimisation on metrics based on the FDC could be, however, a significant alternative (Krause et al. ). For calibration we used 1 month of the available rainfall/runoff data (October/ November 2000), to keep the computational requirements low. The period was selected for data quality (no gaps) and for the variety of rainfall events that occurred. Validation was performed calculating the Nash criterion for each season of the remaining data (nine periods of 4 months without long gaps were identified), in order to

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verify that the choice of a single month, in autumn, for calibration, was not reducing the performances in other periods of the year.

Results and discussion

Scenarios

The calibration procedure provided a high value of the Nash criterion (0.91), meaning a good accuracy of the model. Visual veriﬁcation was also satisfactory (Figure S2, supplementary material, available online at http://www.iwaponline.com/ wst/070/206.pdf). In validation, the Nash criterion values for the available data not used in calibration ranged between 0.72 and 0.89 (median: 0.85), without evident seasonal biases.

The scenarios compared are of two types: specific flow-rate regulations, demanding to implement BMPs to store stormwater and release it at a limited specific flow-rate q* (l/ (s·ha)), and volume regulations, demanding to store and infiltrate a given volume i* (mm) of stormwater. These regulations are the most common in France, and are used in other countries (Faulkner ; Balascio & Lucas ). For more details on their modelling, see Petrucci et al. (). In this example, we consider regulations with several values of specific flow-rate and volume chosen according to French current practices and applied systematically all over the catchment. However, the same procedure can be applied to more complex scenarios, with partial applications of source control. Specific flow-rate regulations were modelled by installing a reservoir downstream of the impervious areas of each subcatchment. Each reservoir has a volume and an outfall sized as a function of q*, according to regulatory local sizing procedure. Volume regulations are modelled as a pervious filter strip downstream of each impervious area. The storage of the strip is sized as a function of i*. Each scenario was simulated on a 10-year period with a 5-minute time-step in order to compute median annual FDCs. We also simulated a synthetic rainfall having a return period of 10 years (triangular, 1-hour duration), in order to compare a classical peak flow-rate indicator to FDC analysis.

Figure 2

Calibration and validation

Scenario simulation In Figure 2 are plotted the median annual FDCs for flow-rate and volume regulations. The reference value (black line) is the FDC calculated by the model with no regulation applied. As discussed before, starting from time-series, it is possible to compute confidence intervals and other statistics on FDC. In this example, we limit to the median FDCs to keep a good readability of results. The two sets of FDCs show immediately a significant difference of behaviours between the two types of source control regulations: while specific flow-rate regulations reduce the frequencies of high flow-rates but increase those of small ones, volume regulations systematically reduce frequencies. When specific flow-rate regulations become stricter (i.e. smaller q*), highest flow-rates progressively disappear, but low flow-rates are increasingly common, exceeding the frequencies of the reference case. The effect of specific flowrate regulations can be summarised by a flattening of the hydrological behaviour of the catchment. The atypical behaviour of the q* ¼ 0.5 l/(s·ha) curve is because of the sizing procedure, which uses rainfall statistics (maximal

Median annual FDCs for ﬂow-rate (left) and volume (right) regulations. FDCs are computed over 10 years of simulation with a 5-minute time-step.

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Q10% (left) and Q10years (right).

intensities) for durations up to 4 days. This is not adapted for long emptying times of a reservoir: with so low an outlet flow, emptying can require 1 week, causing spills for subsequent rain events. This sizing problem cannot be identified by single-event simulations, but easily appears with long-term simulations. Volume regulations, in contrast, reduce flow-rates for each frequency. Small values of i* significantly reduce high-frequency flow-rates but are poorly effective at reducing low-frequency ones. Stricter regulations (higher i*) are able to significantly reduce both. If the comparison of FDCs allows the identifying of global trends, it is possible to show that FDC-based indicators can help decision-making. In Figure 3 are plotted the indicators Q10years, corresponding to the peak flow-rate for a 10-year design rainfall (26.8 mm in 1 hour), and Q10%, representing the flow-rate exceeded 10% of the year (36.5 days). The first graph, representing a classical indicator for stormwater management, shows a linear reduction of peak flow-rates decreasing q*. Because classical stormwater management practices aim to reduce peak flow-rate indicators in order to prevent sewer overflows, this reduction justifies a general trend, among local authorities, to progressively reduce the q* value in regulations: the stricter the rule, the smaller the flow-rates to be managed downstream, and the fewer the sewer overflows and other nuisances. The second graph, issued from the frequency analysis of long-term simulations, presents a different behaviour of flow-rate regulations: the indicator has a significant increase for low values of q*. If there was a constraint on downstream flow-rates, like an erosive threshold for a downstream creek or the activation threshold for a CSO estimated at 1 l/(s·ha) (about 200 l/s), the graph could show that regulations ranging from q* ¼ 5 l/(s·ha) to q* ¼ 1 l/(s·ha) (quite common in France) will significantly worsen downstream conditions.

More generally, this example shows that protecting the environment downstream of an urban area can be an objective contrasting with the protection of the urban area alone. Volume regulations seem able to attain these two objectives simultaneously, but the point here is that similar observations are impossible if only single-event, peak flow-rate analyses are done. Taking into account downstream water bodies through FDCs can improve their protection and cast doubts on the current practices of source control regulations. As the example shows, FDCs are not alternative but complementary to classical peak flow-rate and other analyses. On one side, the concern for water bodies’ protection is growing, but the first purpose of urban stormwater systems remains the protection of people and goods from flooding and other nuisances. FDC analysis, in this context, is just a tool to facilitate the search for a better compromise between different objectives. On the other side, FDCs can provide information on low flows, but they do not preserve temporal patterns: environmental processes depending on the stability of a given level of flow or on high/low flow successions cannot be described by FDC analysis (Sánchez Navarro et al. ).

CONCLUSIONS In this paper, we presented the value of analysing urban stormwater management by the means of FDCs. An analysis of the current modelling practices in urban hydrology and of the literature on the impacts of urban areas on the downstream environment demonstrated that current approaches based on peak ﬂow-rates are not satisfactory. They are not able to inform decision-makers of the impact of their choices on the environment. In this context, we argued that FDCs are a promising tool: they provide a good insight on the impacts on downstream water bodies without

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requiring significant changes in the way urban water studies are done today. FDCs, rarely used today in urban hydrology, represent a practical approach to improve the protection of water bodies. To illustrate the use of FDCs, we presented an example of application to source control regulations in an urban catchment in France. The case-study demonstrates how FDC analysis can modify the conclusions on the impact of different types of regulation: while a typical peak-flows analysis would conclude on the appropriateness of flowrate regulations, the FDCs show their potential negative impacts on the environment, suggesting that volume regulations can be more appropriate. Beside source control, FDCs can be used for almost any situation where flow-rate modelling can be applied, including decisions on new infrastructures, new operation rules or new urban developments. We discussed in general the advantage and requirements of the use of FDCs in urban hydrology, but it should be remembered that information provided by FDCs has to be interpreted according to the specific context. A wise use of FDCs requires an understanding of which changes in the hydrological behaviour are more harmful (or positive) to the specific downstream environment. Local knowledge on water bodies remains necessary to better manage them.

ACKNOWLEDGEMENT We thank Nantes Métropole for providing geographical data about the catchment.

REFERENCES Ashley, R. & Hopkinson, P.  Sewer systems and performance indicators into 21st century. Urban Water 4 (2), 123–135. Balascio, C. & Lucas, W.  A survey of storm-water management water quality regulations in four Mid-Atlantic States. Journal of Environmental Management 90 (1), 1–7. Bertrand-Krajewski, J. L.  Stormwater pollutant loads modelling: epistemological aspects and case studies on the influence of field data sets on calibration and verification. Water Science and Technology 55, 1–17. Booth, D. & Jackson, C.  Urbanization of aquatic systems: degradation thresholds, stormwater detection, and the limits of mitigation. Journal of the American Water Resources Association 33 (5), 1077–1090. Booth, D., Karr, J., Schauman, S., Konrad, C., Morley, S., Larson, M. & Burges, S.  Reviving urban streams: land use,

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hydrology, biology, and human behavior. Journal of the American Water Resources Association 40 (5), 1351–1364. Braud, I., Breil, P., Thollet, F., Lagouy, M., Branger, F., Jacqueminet, C., Kermadi, S. & Michel, K.  Evidence of the impact of urbanization on the hydrological regime of a medium-sized periurban catchment in France. Journal of Hydrology 485, 5–23. Elliott, A. H. & Trowsdale, S. A.  A review of models for low impact urban stormwater drainage. Environmental Modelling & Software 22, 394–405. Faulkner, B.  The control of surface water runoff from new development – UK national ‘policy’ in need of review? Urban Water 1 (3), 207–215. Fennessey, L., Hamlett, J., Aron, G. & LaSota, D.  Changes in runoff due to stormwater management pond regulations. Journal of Hydrologic Engineering 6 (4), 317–327. Gironàs, J., Roesner, L., Davis, J. & Rossman, L.  Storm Water Management Model Applications Manual. United States Environment Protection Agency, Cincinnati, OH, USA. Gupta, H. V., Beven, K. J. & Wagener, T.  Model calibration and uncertainty estimation. In: Encyclopedia of Hydrological Sciences (M. G. Andersen, ed.). Wiley, New York, pp. 2015– 2031. Hunt, W. F. & Tillinghast, E. D.  Relating stormwater control measure (SCM) design standards to stream erosion in Piedmont North Carolina: case studies in Raleigh and Chapel Hill, North Carolina. In: Proceedings of the 12th ICUD, 11–16 September 2011, Porto Alegre, RS, Brazil. Jung, S., Arnaud, F., Bonté, P., Chebbo, G., Lorgeoux, C., Winiarski, T. & Tassin, B.  Temporal evolution of urban wet weather pollution: analysis of PCB and PAH in sediment cores from Lake Bourget, France. Water Science and Technology 57 (10), 1503–1510. Krause, P., Boyle, D. P. & Bäse, F.  Comparison of different efficiency criteria for hydrological model assessment. Advances in Geosciences 5, 89–97. Petrucci, G.  La diffusion du contrôle à la source des eaux pluviales urbaines. Confrontation des pratiques à la rationalité hydrologique. PhD Thesis, Université Paris-Est, Paris, France. Petrucci, G., Rioust, E., Deroubaix, J.-F. & Tassin, B.  Do stormwater source control policies deliver the right hydrologic outcomes? Journal of Hydrology 485, 188–200. Rodriguez, F., Andrieu, H. & Morena, F.  A distributed hydrological model for urbanized areas – model development and applications to case studies. Journal of Hydrology 351 (3–4), 268–287. Rohrer, C., Roesner, L., Mikkelsen, P., Vollertsen, J., HvitvedJacobsen, T. & Ledin, A.  Matching the critical portion of the flow duration curve to minimise changes in modelled excess shear. Water Science and Technology 54 (6–7), 347–354. Roy, A. H., Freeman, M. C., Freeman, B. J., Wenger, S. J., Ensign, W. E. & Meyer, J. L.  Investigating hydrologic alteration as a mechanism of fish assemblage shifts in urbanizing streams. Journal of the North American Benthological Society 24 (3), 656–678. Sánchez Navarro, R., Stewardson, M., Breil, P., de Jalón, D. G. & Eisele, M.  Hydrological impacts affecting endangered

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ﬁsh species: a Spanish case study. River Res. Appl. 23, 511–523. Savic, D. & Khu, S.  Evolutionary computing in hydrological sciences. In: Encyclopedia of Hydrological Sciences, vol. 1 (M. G. Andersen, ed.). Wiley, New York, pp. 331–348. Schütze, M., Campisano, A., Colas, H., Schilling, W. & Vanrolleghem, P. A.  Real time control of urban wastewater systems where do we stand today? Journal of Hydrology 299 (3–4), 335–348. Silva, T., Petrucci, G., Vinçon-Leite, B., Lemaire, B., Garnier, R., Tran Khac, V., Seidl, M., Tassin, B. & Nascimento, N.  Coupling cyanobacteria dynamics and urban runoff modelling: an integrated approach for a tropical lake in Brazil. In: Novatech 2013, 24–27 June 2013, Lyon, France.

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Vogel, R. M. & Fennessey, N. M.  Flow-duration curves. I: New interpretation and conﬁdence intervals. Journal of the American Water Resources Association 120 (4), 485–504. Vogel, R. M. & Fennessey, N. M.  Flow-duration curves. II: A review of applications in water resources planning. Journal of the American Water Resources Association 31 (6), 1029–1039. Walsh, C., Fletcher, T. & Ladson, A. a Stream restoration in urban catchments through redesigning stormwater systems: looking to the catchment to save the stream. Journal of the North American Benthological Society 24 (3), 690–705. Walsh, C., Roy, A., Feminella, J., Cottingham, P., Groffman, P. & Morgan, R. b The urban stream syndrome: current knowledge and the search for a cure. Journal of the North American Benthological Society 24 (3), 706–723.

First received 12 September 2013; accepted in revised form 15 April 2014. Available online 29 April 2014

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Biosorption of uranium(VI) from aqueous solution by biomass of brown algae Laminaria japonica K. Y. Lee, K. W. Kim, Y. J. Baek, D. Y. Chung, E. H. Lee, S. Y. Lee and J. K. Moon

ABSTRACT The uranium(VI) adsorption efficiency of non-living biomass of brown algae was evaluated in various adsorption experimental conditions. Several different sizes of biomass were prepared using pretreatment and surface-modification steps. The kinetics of uranium uptake were mainly dependent on the particle size of the prepared Laminaria japonica biosorbent. The optimal particle size, contact time, and injection amount for the stable operation of the wastewater treatment process were determined. Spectroscopic analyses showed that uranium was adsorbed in the porous inside structure of the biosorbent. The ionic diffusivity in the biomass was the dominant rate-limiting factor; therefore, the adsorption rate was significantly increased with decrease of particle size. From the results of comparative experiments using the biosorbents and other chemical adsorbents/ precipitants, such as activated carbons, zeolites, and limes, it was demonstrated that the brown algae biosorbent could replace the conventional chemicals for uranium removal. As a post-treatment for the final solid waste reduction, the ignition treatment could significantly reduce the weight of

K. Y. Lee (corresponding author) K. W. Kim Y. J. Baek D. Y. Chung E. H. Lee J. K. Moon Korea Atomic Energy Research Institute (KAERI), 989-111 Daedeok-daero, Yuseong-gu, Daejeon, 305-353, Republic of Korea E-mail: lky@kaeri.re.kr S. Y. Lee Department of Mechanical Engineering, Sogang University, 1 Shinsu-dong, Mapo-gu, Seoul, 121-742, Republic of Korea

waste biosorbents. In conclusion, the brown algae biosorbent is shown to be a favorable adsorbent for uranium(VI) removal from radioactive wastewater. Key words

| biosorption, brown algae, Laminaria japonica, radioactivity, uranium, waste reduction

INTRODUCTION Radioactive wastewater containing uranium is mainly generated during the operation of nuclear power plants or related facilities such as nuclear research laboratories and nuclear fuel manufacturing plants. Due to its radioactivity and high toxicity, uranium is one of the most seriously hazardous materials for the environment (Yang & Volesky ), and it is a carcinogen that causes cancer in the human body (Khani ). Uranium occurs naturally in diverse valence states, but it exists mainly as soluble uranium (VI), uranyl ion (UO2þ 2 ), in acidic conditions in the solution (Kim et al. ; Khani ). Some methods have been developed for the treatment of uranium wastewater, including chemical precipitation, adsorption, ion exchange, solvent extraction, evaporation, and electrochemical treatment (Bayramoglu et al. ; Fu & Wang ). However, several drawbacks such as the high treatment cost, secondary waste, and low efﬁciency at low uranium concentration remain unresolved; therefore, the development of an efﬁcient technology for the treatment of uranium-containing wastewater is required. doi: 10.2166/wst.2014.205

Recently, biosorption technology using microorganisms, plants, and other various biomasses has been considered as an alternative to the metal adsorption process using chemical adsorbents (Volesky & Holan ). Studies on uranium biosorption by several species of brown algae such as Sargassum ﬂuitans, Cystoseira indica, Padina sp., Sargassum hemiphyllum, and Undaria pinnatiﬁda have been carried out over the last two decades (Yang & Volesky ; Sakamoto et al. ; Khani et al. ; Khani ). Alginic acid occurring in all brown algae has been considered as a material capable of forming complexes with uranium (Gok & Aytas ) when the non-living biomass of brown algae is used without any pretreatment. The biosorption of uranium by protonated brown algae was studied as an ion exchange process between protons in biomass and soluble uranium ions (Yang & Volesky ). Kinetic, thermodynamic, and equilibrium studies have also been carried out, some of which have shown that the Langmuir isotherm model is suitable for describing the biosorption equilibrium

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of uranium by brown algae (Khani et al. ; Khani ). In order to apply the brown algae biosorbent in the actual uranium removal process, such theoretical studies and comparative evaluations with other materials should be combined. Also, further enhancement schemes are still required to improve the performance of uranium removal from radioactive wastewater and to overcome the drawbacks of former treatment processes. In this study, a feasibility test of biosorption for uranium removal was performed by comparing with former adsorbents and precipitants, for the ﬁrst time. Also, the adsorption properties of uranium by pretreated brown algae were evaluated in different operating conditions.

MATERIALS AND METHODS Uranium wastewater The simulated radioactive wastewater was prepared by dissolving uranyl nitrate (UO2(NO3)2·6H2O) in deionized (DI) water as 10 and 100 mg/l of initial uranium concentrations. In order to evaluate the removal efficiency of uranium at different conditions, the ionic strength of the simulated wastewater was changed by sodium nitrate (NaNO3), and its pH was adjusted to 0 or to 12 by adding nitric acid (HNO3) or sodium hydroxide (NaOH) to the solution. Preparation of adsorbents and precipitants The non-living biomass of brown algae was obtained from the Korean southern coast. Three types of brown algae, Sargassum fulvellum, Undaria pinnatifida, and Laminaria japonica, were selected for the preparation of the surfacemodified biosorbents. The brown algae were dried in an oven at 60 C overnight, and ground in a mortar. They were classified by sieving into four different sizes (>2, 2– 0.2, 0.2–0.05, and <0.05 mm). Surface modification of each classified biomass was carried out by mixing 0.1 M hydrochloric acid (HCl) as 10 g biomass per litre and stirring for 12 h at room temperature. After the acidification, the biomass was rinsed several times with deionized water. The surface-modified biomass was then dried in an oven at 60 C overnight. The uranium adsorption efficiency of several chemical adsorbents and precipitants was compared with that of the prepared biosorbents. The selected adsorbents, activated carbon (AC), acidified-activated carbon (AAC), zeolite types 4A, X and Y, and mordenite (MOR) were tested in W

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the same conditions as those for testing of the biosorbents. AC is a common adsorbent for the removal of various toxic materials, and AAC is known to be more efﬁcient for uranium adsorption because of the increase of surface functional groups (Abbasi & Streat ). In the case of zeolites, the sorption of uranium ions from the liquid phase has been studied using zeolite X, zeolite Y, MOR, and clinoptilolite etc. (Olguin et al. ). It is known that uranyl ions (UO2þ 2 ) are mainly exchanged with the cations from the large cavities of these zeolites. The same procedure was used for the acidiﬁcation of activated carbon. Calcium oxide (CaO), calcium hydroxide (Ca(OH)2), and sodium hydroxide (NaOH) as precipitants were also used for the comparative uranium removal experiments. Except for the AAC, all the other adsorbents and precipitants were commercially available chemicals at a reagent grade.

Adsorption experiments The uranium biosorption was evaluated in various experimental conditions using several types of biosorbents prepared with different particle sizes. All the batch experiments were conducted using a rotary shaker at a room temperature of 25 ± 1 C. The uranium concentrations of the liquid samples obtained from the adsorption experiments were analyzed using a colorimetric method with Arsenazo III (Strelow et al. ) to calculate uranium adsorption capacity (qu, mg/g), distribution coefﬁcient (Kd), and removal efﬁciency (R, %), which are based on the following equations: W

qu ¼ (Ci Cf )V=m

(1)

Kd ¼ (Ci Cf )V=Cf m

(2)

R ¼ (Ci Cf )100=Ci

(3)

where Ci is the initial uranium concentration (mg/l), Cf is the ﬁnal uranium concentration (mg/l) from the supernatant solution, V is the solution volume (ml), and m is the mass of adsorbents (g). The analytical data were assessed using a quality control system, including duplicates, blanks, and standard reference samples, during all procedures for accuracy and precision, which demonstrated percentage error (%) of less than 6.2% and relative standard deviations (%) of less than 4.3%. After the biosorption experiments, the ﬁnally obtained biomass samples contaminated with uranium were analyzed using scanning electron microscopy (SEM) (SNE 4500M,

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RESULTS AND DISCUSSION Comparison of adsorbents

2014

Figure 1 shows that the kinetics of uranium uptake were mainly dependent on the particle size of the prepared L. japonica biosorbent, which gave us the optimal particle size for stable operation of the wastewater treatment process. Although the maximum uranium adsorption capacities did not differ for the particle sizes, the adsorption rate signiﬁcantly increased with decrease of particle size

Uranium adsorption capacity (qu, mg/g) and distribution coefﬁcient (Kd) in DI water containing 100 mg/l uranium (pH 4) after shaking for 6 h with 1 g adsorbents per litre

Activated carbon

Zeolite

AAC

3.5

37.3

3.6 × 10

Figure 1

Adsorption properties of biosorbents

The uranium removal efﬁciencies of several chemical adsorbents and prepared biosorbents were evaluated at 100 mg/l initial uranium concentration and pH 4. One gram per litre of each adsorbent was used in the 6-h batch experiment. Although the chemical adsorbents are commonly considered as useful materials for uranium removal processes, the brown algae biosorbents prepared in this study showed much higher uranium adsorption capacities than the chemical adsorbents compared under the same experimental

Table 1

70.1

condition (Table 1). Also, the result that the distribution coefficients of the brown algae biosorbents were more than 104 strongly demonstrated their high performance for the uranium removal from the solution. Although the data of uranium adsorption capacities were not the maximum values because they are strongly dependent on the initial uranium concentration, the overall removal efficiencies based on the distribution coefficients were similar to or higher than the corresponding literature data (Yang & Volesky ; Khani ). The highest uranium adsorption capacity and distribution coefficient among the biosorbents in this study were obtained with L. japonica biosorbent; further specific evaluations of biosorption were therefore carried out using L. japonica biosorbent.

SEC, Korea) and electron dispersive spectroscopy (EDS) (QUANTAX, Bruker, Germany) for the morphologic and elemental analyses, respectively. In order to measure the weight reduction rate of the biosorbents by ignition treatment, a thermal analyzer (TA) (DSC2920-TGA2950, TA Instruments, USA) was used. The temperature in the TA instrument was increased from 25 C to around 1,100 C at a heating rate of 8 C/min. The weight reduction of MOR was also measured as the control sample. W

6.0 × 10

Brown algae X

26.9 2

37.6 2

3.7 × 10

MOR

70.7 2

6.0 × 10

2.4 × 10

71.6 3

2.5 × 10

S. fulvellum

U. pinnatiﬁda

L. japonica

93.2

94.1

96.4

1.4 × 10

1.6 × 10

2.7 × 104

Uranium removal efficiencies (%) of L. japonica biosorbents (1 g/l) with time in (a) DI water and (b) 1 M NaNO3 containing 100 mg/l uranium at pH 4. (Inset box: enlarged figure of uranium removal efficiencies (%) within 6 h.)

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Figure 2

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Uranium(VI) biosorption by L. japonica biosorbents

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Uranium removal efﬁciencies (%) of L. japonica biosorbents at concentration of (a) 1 g/l and (b) 5 g/l with time in 1 M NaNO3 containing 10 mg/l uranium at pH 4. (Inset box: uranium concentrations (mg/l) in the solution treated with different amounts of biosorbent.)

(inset box of Figure 1(a)). The adsorption rate constants of sizes >2, 2–0.2, 0.2–0.05, and <0.05 mm calculated by using the pseudo second order reaction model were 7.5 × 10 5, 1.5 × 10 3, 2.5 × 10 2 and 2.8 × 10 1 within 6 h, respectively. Moreover, the desorption rate after a long contact time was significantly increased for the small particles. Such a desorption reaction indicates that the uranium ions’ unstable binding is more dominant than the strong chemical complexation onto the biosorbent, although they are exchanged with protons introduced to the binding sites by the acid pretreatment. Particle sizes larger than 0.2 mm showed relatively slow adsorption reaction, and particle sizes smaller than 0.05 mm showed excessive desorption rate, which gave us the optimal range of particle size, 0.2–0.05 mm. The uranium adsorption capacity decreased with the increase of the ionic strength of the solution (Figure 1(b)), which indicates the competitive adsorption with sodium ions. Because the biosorbent does not have selectivity for a specific single ion, it should be considered that treating the actual wastewater containing complex and various ions might have more limitations than the simple matrix solution used in the laboratory. Similar trends for the adsorption kinetics for different particle sizes were observed at the condition of lower initial uranium concentration (10 mg/l) with 1 M sodium nitrate (Figure 2(a)). Because of the high ionic strength in the solution, the residual uranium concentration of the treated solution could not meet the criterion (IAEA recommended criterion of 1 × 10 2 Bq/ml as radioactivity, which is about 1 mg/l of uranium as chemical concentration). The inset box of Figure 2(b) shows that 5 g L. japonica biomass injection in one litre was sufficient for uranium removal to less than 1 mg/l. Also, 2 h of reaction time was enough for the L. japonica biosorbents of these particle sizes (Figure 2(b)).

Although there have been many studies supporting the feasibility of biosorption, we have shown here the most appropriate operating conditions with respect to the biosorbent particle size and sorption kinetics, for the first time. The effect of pH on the uranium adsorption efficiency of L. japonica biosorbent showed that the highest uptake occurred at pH 4 in a solution with 1 M sodium nitrate (Table 2), which is consistent with previous uranium biosorption studies (Khani et al. ; Gok & Aytas ). While uranium exists mainly as soluble uranyl ion (UO2þ 2 ) in strongly acidic conditions, other hydrolyzed complex ions þ 2þ ((UO2)3(OH)þ 5 , (UO2)2(OH)2 , and UO2OH ) coexist at pH 4, which is a more favorable condition for adsorption onto the binding sites (Yang & Volesky ; Kim et al. ). On the other hand, anionic complex ions are predominant in neutral and alkaline conditions with low uranium contents (Kim et al. ), which causes a decrease of uranium removal efficiency with a pH increase to more than 4. Figure 3 shows SEM images of L. japonica biosorbents. It was clearly observed that the inside of L. japonica biosorbent was a porous structure, whereas the outer skin was not. Such a porous inside structure might enable the uranyl and other hydrolyzed ions to easily infiltrate and be adsorbed. Figure 4 presents a SEM image of the cross-section of

Table 2

Uranium removal efﬁciency (%) of L. japonica biosorbents (0.2–0.05 mm, 5 g biosorbents per litre) in 1 M NaNO3 containing 100 and 10 mg/l uranium after shaking for 6 h at different pHs

6 a

U 100 mg/l

13.1

71.8

85.7

87.9

n.a.

U 10 mg/l

36.3

85.6

89.2

94.7

92.0

90.0

87.1

100 mg/l

initial

uranium

Not

analyzed.

concentration.)

(Hydroxide

precipitation

occurred

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Figure 3

SEM images of L. japonica biosorbents.

Figure 4

(a) SEM image, (b) quantitative analysis, and (c) elemental mapping by EDS of cross-sectional L. japonica biosorbent after uranium adsorption in 1 M NaNO3.

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Uranium(VI) biosorption by L. japonica biosorbents

L. japonica biosorbent with semi-quantitative analysis and elemental mapping by EDS after uranium adsorption in 1 M sodium nitrate condition. Uranium was well distributed in the biosorbent, which could explain the different uranium adsorption properties for the different particle sizes. As mentioned above, the maximum uranium adsorption capacities did not differ for the particle sizes because the total binding sites were the same. However, the ionic diffusivity in the biomass was the dominant rate-limiting factor; therefore, the adsorption rate was significantly increased with a decrease of particle size. A considerable amount of sodium was confirmed from both semi-quantitative analysis and elemental mapping, whereas the result of uranium adsorption in DI water was not confirmed (data not shown). The competitive adsorption of uranium and sodium ions was visibly confirmed from these results. Performance evaluation of biosorbents Up to the present, the chemical precipitation of heavy metals from wastewater has been considered as the most traditional and applicable process due to its simplicity and inexpensive cost (Fu & Wang ). Also, for the uranium wastewater treatment in nuclear industries, such as nuclear fuel manufacturing plants, chemical precipitation using limes (CaO and/or Ca(OH)2) has been one of the major processes. Figure 5 shows the uranium removal efficiencies of L. japonica biosorbents and other precipitants (5 g/l) in 1 M sodium nitrate containing different uranium concentrations. For the removal of a relatively high concentration of uranium, the limes precipitated the uranium with a long operation time (more than 1 week) in both shaking and static conditions. The adsorption reaction of L. japonica biosorbents was extremely fast, so

Figure 5

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that the maximum removal of uranium was achieved within 2 h. In general, chemical precipitation is adapted to treat wastewater containing high concentrations of metal ions and it is ineffective when the metal ion concentration is low, because the chemical precipitation is theoretically impossible when the concentration is lower than the solubility of the precipitates. For example, uranium is precipitated as uranium peroxide (UO4) by adding excess hydrogen peroxide (H2O2) to an acidic solution of uranyl ions (Kim et al. ). Although this technique has been considered an effective method for uranium recovery, the solution after the uranium peroxide precipitation still contains several mg/l of residual uranium, which should be removed for the wastewater release. In the case of the lower initial uranium concentration (1 mg/l) simulating the above cases, the precipitants did not remove the uranium; however, the biosorbents reduced the uranium concentration to a non-detectable level in less than 30 min (Figure 5(b)). These comparative experiments strongly demonstrated the satisfactory performance of the prepared biosorbents compared with others. Post-treatment for solid waste reduction The management of radioactive solid waste is very important due to the extremely high cost of disposal. Therefore, the reduction of radioactive solid waste has been one of the technologies urgently required in the nuclear industry. This section ﬁnally suggests the possibility of volume reduction of waste biosorbents after the uranium removal from wastewater. Because the major component of brown algae is organic carbon, it should be decomposed at a high temperature. Figure 6(a) shows the weight reduction of L. japonica biosorbent and MOR (as a control) by ignition treatment. Although

Uranium removal efﬁciencies (%) of L. japonica biosorbents and other precipitants (CaO, Ca(OH)2, and NaOH) at 5 g/l in 1 M NaNO3 containing (a) 10 mg/l and (b) 1 mg/l uranium at pH 4. (Solid lines: shaking condition, dotted lines: static condition.)

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Figure 6

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Uranium(VI) biosorption by L. japonica biosorbents

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(a) Weight (%) reduction of MOR and L. japonica biosorbent by ignition treatment (heating rate: 8 C/min, inset box: loss on ignition (%) after heating for 6 h at 550 C). (b) Picture of L. japonica biosorbent before and after ignition treatment. W

the reduction of MOR, an alumino-silicate mineral, was negligible, that of the biosorbent was significant. The inset box of Figure 6(a) shows that the loss on the ignition of biosorbent was 75.2% after heating for 6 h at 550 C. Also, the significant volume reduction could be visibly confirmed by comparison of before and after ignition treatment (Figure 6(b)). Therefore, the brown algae biosorbents have a strong advantage when reduced by ignition treatment.

government (MSIP) (No. NRF-2012M2A8A5025658 & 2012M2A8A4055325).

CONCLUSIONS From the results of comparative experiments using biosorbents and other chemical adsorbents/precipitants, it was demonstrated that the brown algae biosorbent could replace the conventional chemicals for uranium removal. We have shown here that the kinetics of uranium uptake was mainly dependent on the particle size of the biosorbents. The optimal particle size, contact time, and injection amount for the most appropriate operation of the wastewater treatment process were determined, for the first time. As a post-treatment for the final solid waste reduction, the ignition treatment could significantly reduce the weight of waste biosorbents. In conclusion, the surface-modified biosorbent from brown algae has been shown to be a favorable adsorbent for uranium removal from radioactive wastewater.

ACKNOWLEDGEMENT This work was supported by the National Research Foundation of Korea (NRF) grant funded by the Korea

REFERENCES Abbasi, W. A. & Streat, M.  Adsorption of uranium from aqueous solutions using activated carbon. Sep. Sci. Technol. 29 (9), 1217–1230. Bayramoglu, G., Celik, G. & Yakup Arica, M.  Studies on accumulation of uranium by fungus Lentinus sajor-caju. J. Hazard. Mater. B136, 345–353. Fu, F. & Wang, Q.  Removal of heavy metal ions from wastewater: a review. J. Environ. Manage. 92, 407–418. Gok, C. & Aytas, S.  Biosorption of uranium (VI) from aqueous solution using calcium alginate beads. J. Hazard. Mater. 168, 369–375. Khani, M. H.  Uranium biosorption by Padina sp. algae biomass: kinetics and thermodynamics. Environ. Sci. Pollut. Res. 18, 1593–1605. Khani, M. H., Keshtkar, A. R., Ghannadi, M. & Pahlavanzadeh, H.  Equilibrium, kinetic and thermodynamic study of the biosorption of uranium onto Cystoseira indica algae. J. Hazard. Mater. 150, 612–618. Kim, K. W., Kim, Y. H., Lee, S. Y., Lee, J. W., Joe, K. S., Lee, E. H., Kim, J. S., Song, K. & Song, K. C.  Precipitation characteristics of uranyl ions at different pHs depending on the presence of carbonate ions and hydrogen peroxide. Environ. Sci. Technol. 43, 2355–2361. Kim, K. W., Hyun, J. T., Lee, K. Y., Lee, E. H., Lee, K. W., Song, K. C. & Moon, J. K.  Effects of the different conditions of uranyl and hydrogen peroxide solutions on the behaviour of the uranium peroxide precipitation. J. Hazard. Mater. 193, 52–58. Olguin, M. T., Solache-Rios, M., Acosta, D., Bosch, P. & Bulbulian, S.  Uranium sorption in zeolite X: the valence effect. Micro. Mes. Mater. 28, 377–385.

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Sakamoto, N., Kano, N. & Imaizumi, H.  Biosorption of uranium and rare earth elements using biomass of algae. Bioinorg. Chem. Appl. art. 706240, 8 pp. http://dx.doi.org/ 10.1155/2008/706240. Strelow, F. W., KoKot, M. L., Walt, T. N. V. D. & Bhaga, B.  Rationalized determination of uranium in rocks for geochemical prospecting using separation by ion

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exchange chromatography and spectrophotometry with arsenazo III. J. South African Chem. Inst. XXIX, 97–104. Volesky, B. & Holan, Z. R.  Biosorption of heavy metals. Biotechnol. Prog. 11 (3), 235–250. Yang, J. & Volesky, B.  Biosorption of uranium on Sargassum biomass. Water Res. 33 (15), 3357–3363.

First received 6 November 2013; accepted in revised form 14 April 2014. Available online 26 April 2014

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Combination of ion exchange and partial nitritation/ Anammox process for ammonium removal from mainstream municipal wastewater Andriy Malovanyy, Elzbieta Plaza, Jozef Trela and Myroslav Malovanyy

ABSTRACT In this study, a new technology of nitrogen removal from mainstream municipal wastewater is proposed. It is based on ammonium removal by ion exchange and regeneration of ion exchange material with 10–30 g/L NaCl solution with further nitrogen removal from spent regenerant by partial nitritation/Anammox process. Inﬂuence of regenerant strength on performance of ion exchange and biological parts of the proposed technology was evaluated. Moreover, the technology was tested in batch mode using pretreated municipal wastewater, strong acid cation (SAC) resin and partial nitritation/Anammox biomass. It was shown that with ion exchange it is possible to remove 99.9% of ammonium from wastewater while increasing the concentration of ammonium in spent regenerant by 18 times. Up to 95% of nitrogen from spent regenerant, produced by regeneration of SAC resin with 10 g/L NaCl solution, was removed biologically by partial nitritation/Anammox biomass. Moreover, the possibilities of integration of the technology into municipal wastewater treatment

Andriy Malovanyy (corresponding author) Elzbieta Plaza Jozef Trela Department of Sustainable Technology, Environmental Science and Engineering, Royal Institute of Technology (KTH), Teknikringen 76, 10044, Stockholm, Sweden E-mail: andriym@kth.se Andriy Malovanyy Myroslav Malovanyy Department of Industrial Ecology and Sustainable Environmental Management, Lviv Polytechnic National University, Bandery 12, 79013 Lviv, Ukraine

technology, and the challenges and advantages are discussed. Key words

| ammonium, Anammox, ion exchange, partial nitritation, strong acid cation resin

INTRODUCTION The Anammox process is mainly applied for treatment of highly concentrated ammonium streams, such as anaerobic digestion reject water, and the number of full-scale plants is constantly growing. Currently, there is a big interest in application of the Anammox process in nitrogen removal from the main stream of municipal wastewater. Since no organic carbon is consumed during the process more biogas can be produced from the organic content of the wastewater. This can be done for example by wastewater treatment in an upﬂow anaerobic sludge blanket (UASB) reactor or removal of organic matter by ﬂoculation with following digestion. However, there are several challenges of partial nitritation/Anammox process:

•

Low yield of Anammox bacteria. According to Strous et al. () yield of Anammox bacteria is 0.052 g/g N. Lower yield together with low inﬂowing nitrogen concentration (around 25–50 mg NH4-N/L) leads to low production of Anammox biomass per 1 m3 of treated wastewater. Therefore, effective retention of Anammox biomass in a reactor is important.

doi: 10.2166/wst.2014.208

•

Low rates. Lower temperature of mainstream wastewater leads to lower rates of nitriﬁcation and Anammox processes. The reported activation energies for nitritation and Anammox processes are similar and are in the ranges of 60–72 kJ/mol (van Hulle et al. ) and 63– 85 kJ/mol (Dosta et al., ; Fernández et al. ) respectively. Therefore, decrease of temperature from 30 C (temperature in reactor treating reject water) to 15 C (temperature of municipal wastewater during winter period) should lead to a decrease of rate by 71– 83%. Moreover, further decrease of rate can be expected because of lower ammonium and nitrite concentrations in the reactor. Nitrite-oxidizing bacteria (NOB) suppression. In the recent studies on application of Anammox process for mainstream municipal wastewater treatment, it was shown that the biggest challenge is controlling competition between ammonium-oxidizing bacteria (AOB) and NOB for oxygen and between NOB and Anammox for nitrite (Wett et al. ). Selection of AOB over NOB is not always successful because of lower activation W

•

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energy of the nitrite oxidation process and consequently higher growth rate of NOB (comparing to AOB) at lower temperatures. Moreover, inhibition of NOB with free ammonia, the mechanism that is often used for NOB suppression in reject water treatment, is not possible because of low ammonium concentration in treated municipal wastewater. A possible solution for overcoming the above-described limitations is a two-step process with concentration of ammonium from municipal wastewater by ion exchange followed by removal of ammonium from concentrated solution by partial nitritation/Anammox process. Ion exchange gives the possibility to selectively concentrate ammonium ions without concentrating un-ionized inorganic and organic substances and suspended material. Several ion exchange materials can be used for ammonium concentration such as natural and synthetic zeolites and strong acid cation (SAC) resin. Application of natural zeolite of clinoptilolite type for ammonium removal from wastewater is extensively studied due to its selective ion exchange properties (Hedström ). SAC resin has higher exchange capacity than natural zeolite and was proposed for ammonium removal in a number of studies (Leakovic´ et al. ; Chen et al. ). After the exchange capacity of ion exchangers is exhausted it can be regenerated by contacting with NaCl solþ ution. However, the NHþ 4 ↔ Na exchange is not complete and the spent regenerant has high residual NaCl content. Salinity has a negative impact on activities of AOB and Anammox bacteria. However, it was shown that with a step-wise increase of NaCl concentration it was possible to adapt bacteria to salinity of 30 g/L (Windey et al. ). The aim of this work was to test the proposed two-step technology of ammonium removal in batch mode and discuss the possibilities and challenges of its application. SAC resin was used to concentrate ammonium from real wastewater with further removal from spent regenerant by partial nitritation/Anammox process. Impact of regenerant strength on performance of the concentration process and biological nitrogen removal was also studied.

MATERIALS AND METHODS Ion exchange experiments Concentration of ammonium was studied using a glass column with diameter of 10 mm ﬁlled with 30.5 mL of SAC resin KU-2-8 in sodium form (sulfonated polystyrene type,

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divinylbenzene content – 8%, grain size – 0.6–1.2 mm, and total static capacity – 1.8 eq/L). In total, eight runs of exhaustion/regeneration were performed using both synthetic (runs 1–3) and municipal (runs 4–8) wastewater. Synthetic wastewater had ion content similar to municipal wastewater – ammonium concentration of 40 mg NH4-N/L, other cations in concentrations as described in Semmens et al. () and pH of 6.2. Municipal wastewater, after organic matter removal in UASB reactor, filtered through a filter with pore size of 1.6 μm was used. After breakthrough detection, the resin was regenerated counter-currently with 10–30 g/L NaCl solution. Biomass Kaldnes rings (K1) with partial nitritation/Anammox biomass were taken from a pilot-scale moving-bed biofilm reactor treating reject water and located at Hammarby Sjöstadsverk research station in Stockholm (Yang et al. ). Based on the results of microscopical characterization of biomass (Winkler et al. ), the main groups of microorganisms in the biofilm were Anammox bacteria and AOB. Activity tests Activities of AOB, NOB, and aerobic heterotrophs were determined by oxygen uptake rate (OUR) tests. The tests were done using methodology based on addition of specific inhibitors of AOB and NOB, which is described in detail in Surmacz-Gorska et al. (). Specific activity of Anammox bacteria (SAA) was determined by the methodology, described in Dapena-Mora et al. (), which is based on gas pressure measurement. Partial nitritation/Anammox batch tests The first 700 mL of spent regenerant, obtained in a preceding ion exchange experiment, was supplemented with 8.4 mg of NaHCO3 per mg of NH4-N, which corresponds to 135% of theoretical alkalinity needed for removal of ammonium through the partial nitritation/Anammox pathway, and a filling volume of 250 mL of biomass carriers was added. Mixing was provided by a magnetic stirrer and dissolved oxygen (DO) was controlled at the level of 1.5 mg/L in the first test and was decreased to 1.0 mg/L in the other two tests. The tests were performed at 22 C. Depending on the course of nitrogen removal, pH changed in the range of 7.4–7.9. Samples were analyzed for ammonium, nitrite, and nitrate using Hach–Lange standard cuvette tests. W

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RESULTS Inﬂuence of NaCl concentration on partial nitritation/ Anammox process performance Inﬂuence of NaCl concentration on nitrogen removal through partial nitritation/Anammox pathway was evaluated separately for the two biological steps – nitritation and Anammox by short-term OUR and SAA tests, respectively, which were done at salinities of medium in the range 0–30 g/L. Results of OUR tests (Figure 1(a)) showed that with increase of salinity the activity of both AOB and heterotrophs decreases. Activity of NOB was on the very low level. Therefore, the impact of salinity on these microorganisms cannot be evaluated. Increase of NaCl concentration to 10 g/L leads to decrease of AOB activity by 20–40% and further increase to 30 g/L causes loss of 70–80% of activity. Results of SAA tests (Figure 1(b)) showed that inhibition of Anammox bacteria is stronger than for AOB. At NaCl concentration of 10 g/L between 25 and 60% of activity in

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unstressed conditions can be expected. Increase of salinity to concentrations over 15 g/L leads to activity decrease to the detection limit of the method. These results are in agreement with previous studies (Windey et al. ; DapenaMora et al. ) where similar inhibitory effects were evaluated. Influence of regenerant strength on performance of ammonium concentration by ion exchange Experiments with different regenerant strength (10, 20, and 30 g/L NaCl solutions) were made aiming to find a balance between effectiveness of the ammonium concentration process and biological nitrogen removal. In these experiments, SAC resin was saturated with ammonium from synthetic wastewater and then regenerated with application of regenerant of different strength. Breakthrough curves of the three runs (Figure 2(a)) follow the same pattern, since wastewater content and flow rate remained the same in the three runs. Ammonium breakthrough capacity (calculated as capacity when Ceff ¼ 0.05Cin) of 0.38 eq/L was reached. Regeneration of resin (Figure 2(b)) with NaCl solutions of different concentrations required different volume of regenerant for complete ammonium recovery. The concentration of ammonium in spent regenerant was 239, 382, and 508 mg/L when 10, 20, and 30 g/L NaCl solutions, respectively, were used as regenerant. Therefore, it was possible to concentrate ammonium from synthetic wastewater with a 6–12.7-fold increase of concentration. Ammonium concentration from municipal wastewater

Figure 1

Inﬂuence of NaCl concentration on activity of aerobic (a) and Anammox (b) bacteria.

In order to confirm the results, obtained with synthetic wastewater, five runs of exhaustion/regeneration were performed, where pretreated municipal wastewater as a source of ammonium was used. In the first two runs regenerant with NaCl concentration of 30 g/L was used, whereas in the latter three runs a less concentrated 10 g/L NaCl solution was used. Breakthrough of ammonium in the five runs with municipal wastewater was detected after passing 4–6.5 L of wastewater (Figure 3(a)), and such a difference was caused by different initial ammonium concentration in separate runs. Run 5 was stopped just after reaching breakthrough, whereas the other runs were allowed to continue until reaching one-quarter of the of initial ammonium concentration. Regeneration of resin after application of municipal wastewater followed the same pattern, and for complete

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Figure 2

Inﬂuence of regenerant strength on performance of ammonium concentration: (a) exhaustion; and (b) regeneration.

Figure 3

Ammonium concentration from municipal wastewater: (a) exhaustion (modiﬁed from Malovanyy et al. 2013); and (b) regeneration.

regeneration it was necessary to supply 0.33 and 0.7 L of regenerant when 30 and 10 g/L regenerant was used, respectively. These results are in agreement with the results of runs 1–3, where synthetic wastewater was used. Results of run 5 demonstrate that it is possible to reach efficiency of ammonium removal of 99.9% and still increase ammonium concentration by 18 times (Table 1). Ammonium removal from spent regenerant using partial nitritation/Anammox process According to results of the short-term activity tests, spent regenerant of runs 4 and 5 had a NaCl content which would totally inhibit AOB and Anammox bacteria. Therefore, biological removal of nitrogen from spent regenerant of these runs was not tested. Nitrogen removal from spent regenerant of runs 6–8 was performed in three separate medium-term batch tests and removal efficiency of up to 95% was reached (Table 2). According to Anammox stoichiometry maximal efficiency of partial nitritation/Anammox

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is 89%, which indicates that part of the nitrogen was removed by heterotrophic denitrifiers. Such efficiencies were also observed in the pilot-scale reactor from which the biomass was taken (Yang et al. ). During the first batch test, DO concentration was kept at 1.5 mg/L, which led to high residual nitrite concentration. Therefore, in the following two batch tests DO concentration was maintained at 1.0 mg/L and no nitrite accumulation was observed. With these tests, it was shown that spent regenerant could be treated with the partial nitritation/Anammox process with high efficiency.

DISCUSSION Ion exchange The results presented above showed a possibility of combining ion exchange with autotrophic nitrogen removal. Concentration of ammonium ions from wastewater was

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Table 1

Ion exchange and partial nitritation/Anammox for ammonium removal

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Results of ion exchange experiments

Concentration in wastewater

Volume of wastewater

Break-through

Regenerant

Volume of regenerant

Average concentration in spent regenerant

Concentration

Run

(mg NH4-N/L)

applied (L)

capacity (eq/L)

(g NaCl/L)

required (L)

(mg NH4-N/L)

increase factor

5.7

0.38

508

12.7

5.8

0.39

0.50

382

9.5

5.7

0.39

0.80

239

26.6

8.4

0.38

0.33

581

21.8

24.8

6.5

0.36

0.34

445

17.9

40.4

5.1

0.41

0.70

367

9.1

21.8

6.0

0.27

0.70

187

8.6

37.9

5.4

0.40

0.70

330

8.7

Table 2

Nitrogen removal from spent regenerant by partial nitritation/Anammox

Final nitrogen concentrations Starting ammonium Run

concentration (mg NH4-N/L)

NH4þ -N (mg/L)

NO 2 -N (mg/L)

NO 3 -N (mg/L)

Removal rate, g N/(m2 d)

Treatment efﬁciency (%)

367

29.05

21.25

0.71

187

0.1

1.06

7.93

0.46

330

3.44

17.02

1.06

possible both when the synthetic wastewater and pretreated municipal wastewater was used. In this study, SAC resin was used as an ion exchange material, which has high exchange capacity and fast regeneration rates. However, it is not selective to ammonium ion and will concentrate also other ions from wastewater. Other ion exchange materials, which are more selective for ammonium, can also be applied. Zeolite rock of clinoptilolite type was used for ammonium removal in a number of studies and reported concentrations of ammonium in spent regenerant were in the range of 100– 280 mg when 17.5–35 g/L solution of NaCl was used for regeneration (Semmens & Porter ; Liberti et al. ). Therefore, with natural zeolites a lower concentration of ammonium can be reached, but the removal of ammonium is more selective. Process schemes There are several possibilities for integration of ammonium removal with ion exchange into the municipal wastewater treatment process conﬁguration. Since organic matter is not retained by ion exchange material in considerable amounts and therefore does not enter the partial nitritation/Anammox

reactor, removal of ammonium can be placed either before or after the organic matter removal step. The first option is to remove ammonium after the steps of primary treatment and high-rate activated sludge (HRAS) process (Figure 4(a)). In order to avoid clogging of the ion exchange column and high content of organics in spent regenerant, it is necessary to have a preceding step of suspended solids removal, for example using sand filters, before the ion exchange step. The disadvantage of this system is that careful control of sludge age is needed in order to limit the nitrification process. Some ammonium still will be converted to nitrite and nitrate and be present in the effluent. Another option is to apply anaerobic digestion of organics (e.g. in UASB reactor) instead of the HRAS step (Figure 4(b)). In this case, all ammonium content from the wastewater can be removed. Alternatively, ammonium can be removed by ion exchange before the organics removal step, which can be either of aerobic type (e.g. activated sludge process) or of anaerobic type (e.g. using UASB reactor) (Figure 4(c)). The disadvantage of this system is that careful control of nitrogen supply to the following step of dissolved organics removal is

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Figure 4

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Ion exchange and partial nitritation/Anammox for ammonium removal

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Possibilities of integration of ammonium removal by ion exchange in municipal wastewater treatment: SCR – screens, GC – grit chamber, PS – primary settler, AB – aeration basin, SS – secondary settler, SF – sand ﬁlter, IE – ion exchange column, PN/A – partial nitritation/Anammox reactor, OR – organics removal, UASB – UASB reactor.

needed. Bacteria need nitrogen as a nutrient for their cell growth; therefore, a low level off ammonium needs to be left in the wastewater after the ion exchange step. If too much ammonium is supplied it will cause lower overall nitrogen removal efﬁciency. Technology limitations and advantages The main challenge of biological treatment of spent regenerant is to adapt biomass to elevated salinities. As mentioned earlier, bacteria can adapt to salinity of 30 g/L when stepwise increase of salinity is provided. However, this adaptation period can be a long-term process and the adaptation strategies need to be further studied. Another limitation is that based on stoichiometry of nitritation and Anammox processes, 1.17 moles of alkalinity needs to be

added per 1 mole of ammonium nitrogen. Moreover, treated spent regenerant can be recycled for ion exchange material regeneration, but part of it needs to be exchanged because of accumulation of other ions in it. Nitrate is inevitably generated due to the Anammox reaction and its removal from the high-concentrated stream needs to be done by denitrification in a separate reactor and supply of external carbon source and/or recharge of regenerant. Despite the limitations of the technology, it offers several advantages. First, very high efficiency of nitrogen removal by ion exchange (up to 99.9% as shown in this study) can be reached. The efficiency of nitrogen removal of the whole system was in the range of 83–91%. Secondly, NOB suppression is reached more easily, because: (1) higher nitrogen concentration allows the use of free ammonia to inhibit NOB; (2) NOB are more sensitive to NaCl (Liu et al.

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); (3) since regenerant is recirculated, it is more feasible to increase the temperature of regenerant, so that AOB growth rate is higher than that of NOB and higher total rates of nitrogen removal could be maintained. Moreover, since ammonium concentration in inflow to the bioreactor is high, comparing to mainstream condition, it is easier to maintain high Anammox biomass concentration. For startup of the process, developed biomass from the reject water treatment plant may be used for inoculation. Comparing to other proposed technologies, where ammonium is concentrated from wastewater by ion exchange and then removed through nitrification/denitrification (Semmens & Porter ) or electrolysis (Shelp & Seed ), the economics of the proposed technology is better, since external carbon addition or high electricity consumption is not required.

CONCLUSIONS Using SAC resin, it is possible to concentrate ammonium from municipal wastewater with increase of ammonium concentration by nine or 18–22 times if NaCl solution with concentrations of 10 or 30 g/L, respectively, is used as regenerant. Inhibition of AOB and Anammox bacteria by NaCl was studied in short-term tests, and it was shown that at NaCl concentration of 10 g/L partial nitritation/Anammox biomass is inhibited only partially. Therefore, non-adapted biomass can be used for removal of ammonium from the spent regenerant of ion exchange. These results were conﬁrmed in mediumterm tests, where partial nitritation/Anammox biomass was used to treat spent regenerant, obtained in previous ion exchange experiments. Based on the obtained results, it can be concluded that it is possible to combine ion exchange with partial nitritation/Anammox process for removal of nitrogen from mainstream municipal wastewater. Using this technology, it is possible to better control NOB suppression. However, external alkalinity need is the major obstacle.

ACKNOWLEDGEMENTS We gratefully acknowledge scholarships provided by Swedish Institute and Ministry of Education and Science of Ukraine to Andriy Malovanyy for his PhD studies. The experimental work was performed at Hammarby Sjöstadsverk (Swedish Water Innovation Center). Financial support for research by Swedish Water Development (SVU) and Swedish Environmental Research Institute (IVL) is greatly appreciated. The authors would like to

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thank Isaac Owusu-Agyeman (MSc student) for his valuable input in the practical work of this research.

REFERENCES Chen, J. P., Chua, M.-L. & Zhang, B.  Effects of competitive ions, humic acid, and pH on removal of ammonium and phosphorous from the synthetic industrial effluent by ion exchange resins. Waste Management 22, 711–719. Dapena-Mora, A., Fernández, I., Campos, J. L., Mosquera-Corral, A., Méndez, R. & Jetten, M. S.  Evaluation of activity and inhibition effects on Anammox process by batch tests based on the nitrogen gas production. Enzyme and Microbial Technology 40, 859–865. Dapena-Mora, A., Vázquez-Padín, J. R., Campos, J. L., MosqueraCorral, A., Jetten, M. S. & Méndez, R.  Monitoring the stability of an Anammox reactor under high salinity conditions. Biochemical Engineering Journal 51, 167–171. Dosta, J., Fernández, I., Vázquez-Padín, J. R., Mosquera-Corral, A., Campos, J. L., Mata-Alvarez, J. & Méndez, R.  Short- and long-term effects of temperature on the Anammox process. Journal of Hazardous Materials 154, 688–693. Fernández, I., Plaza, E., Trela, J., Hultman, B. & Méndez, R.  Evaluation of deammonification process by response surface models. Water, Air, & Soil Pollution 215, 299–309. Hedström, A.  Ion exchange of ammonium in zeolites: a literature review. Journal of Environmental Engineering 127, 673–681. Leakovic´, S., Mijatovic´, I., Cerjan-Stefanovic´, Š. & Hodžic´, E.  Nitrogen removal from fertilizer wastewater by ion exchange. Water Research 34, 185–190. Liberti, L., Boari, G., Petruzzelli, D. & Passino, R.  Nutrient removal and recovery from wastewater by ion exchange. Water Research 15, 337–342. Liu, S., Yang, F., Gong, Z. & Su, Z.  Assessment of the positive effect of salinity on the nitrogen removal performance and microbial composition during the start-up of CANON process. Applied Microbiology and Biotechnology 80, 339–348. Malovanyy, A., Sakalova, H., Plaza, E. & Malovanyy, M.  Concentration of ammonium from municipal wastewater using ion exchange process. Desalination 329, 93–102. Semmens, M. & Porter, P.  Ammonium removal by ion exchange: using biologically restored regenerant. Journal WPCF 51, 2928–2940. Semmens, M., Klieve, J., Schnobrich, D. & Tauxe, G. W.  Modeling ammonium exchange and regeneration on clinoptilolite. Water Research 15, 655–666. Shelp, G. C. & Seed, L. P.  Electrochemical Treatment of Ammonia in Waste-Water. US Patent 7160430 B2, 9 January 2007. Strous, M., Heijnen, J. J. & Kuenen, J. G.  The sequencing batch reactor as a powerful tool for the study of slowly growing anaerobic ammonium-oxidizing microorganisms. Applied Microbiology and Biotechnology 50, 589–596. Surmacz-Gorska, J., Gernaey, K., Demuynck, C., Vanrolleghem, P. & Verstraete, W.  Nitrification monitoring in activated

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sludge by oxygen uptake rate (OUR) measurements. Water Research 1354, 1228–1236. van Hulle, S. W. H., Vandeweyer, H. J. P., Meesschaert, B. D., Vanrolleghem, P. A., Dejans, P. & Dumoulin, A.  Engineering aspects and practical application of autotrophic nitrogen removal from nitrogen rich streams. Chemical Engineering Journal 162, 1–20. Wett, B., Omari, A., Podmirseg, S. M., Han, M., Akintayo, O., Gómez Brandón, M., Murthy, S., Bott, C., Hell, M., Takács, I., Nyhuis, G. & O’Shaughnessy, M.  Going for mainstream deammoniﬁcation from bench to full scale for maximized resource efﬁciency. Water Science and Technology 68, 283–289.

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Windey, K., de Bo, I. & Verstraete, W.  Oxygen-limited autotrophic nitrification-denitrification (OLAND) in a rotating biological contactor treating high-salinity wastewater. Water Research 39, 4512–4520. Winkler, M., Yang, J., Kleerebezem, R., Plaza, E., Trela, J., Hultman, B. & van Loosdrecht, M. C. M.  Nitrate reduction by organotrophic Anammox bacteria in a partial nitrifying granular sludge and a moving bed biofilm reactor. Bioresource Technology 114, 217–223. Yang, J., Trela, J., Plaza, E. & Tjus, K.  N2O emissions from a one stage partial nitrification/Anammox process in moving bed biofilm reactors. Water Science and Technology 68, 144–152.

First received 11 December 2013; accepted in revised form 15 April 2014. Available online 26 April 2014

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Ultrasonic degradation of 1-H-benzotriazole in water Henry Zúñiga-Benítez, Jafar Soltan and Gustavo Peñuela

ABSTRACT This paper reports on the effect of different parameters of ultrasonic power, pollutant initial concentration, pH and the presence of co-existing chemical species (oxygen, nitrogen, ozone, and radical scavengers) on the ultrasonic degradation of the endocrine disruptor 1-H-benzotriazole. Increasing the 1-H-benzotriazole initial concentration from 41.97 to 167.88 μM increased the pollutant degradation rate by 40%. Likewise, a high applied ultrasonic power enhanced the extent of 1-H-benzotriazole removal and its initial degradation rate, which was accelerated in the presence of ozone and oxygen, but inhibited by nitrogen. The most favorable pH for the ultrasonic degradation was acidic media, reaching ∼90% pollutant removal in 2 h. The hydroxyl free radical concentration in the reaction medium was proportional to the ultrasound power and the irradiation time. Kinetic models based on a Langmuir-type mechanism were used to predict the pollutant sonochemical

Henry Zúñiga-Benítez Gustavo Peñuela Grupo GDCON, Facultad de Ingeniería, Sede de Investigación Universitaria (SIU), Universidad de Antioquia, Calle 70 No. 52 -21, Medellín, Colombia Jafar Soltan (corresponding author) Department of Chemical and Biological Engineering, University of Saskatchewan, 57 Campus Drive, Saskatoon, SK S7N 5A9, Canada E-mail: j.soltan@usask.ca

degradation. It was concluded that degradation takes place at both the bubble–liquid interfacial region and in the bulk solution, and OH radicals were the main species responsible for the reaction. Hydroxyl free radicals were generated by water pyrolysis and then diffused into the interfacial region and the bulk solution where most of the solute molecules were present. Key words

| advanced oxidation process, benzotriazole, degradation, ultrasound, water treatment

INTRODUCTION Emerging pollutants and endocrine disruptors (compounds that alter the hormonal balance) are present in various commercial products with wide use in industry and domestic applications (Bolong et al. ). 1-H-benzotriazole (BZ), a polar high production-volume chemical, has been regarded as an emerging pollutant (Xu et al. ). This compound and its derivatives are commonly used in several applications including corrosion inhibitor for metals, antifoggant in photography and UV stabilizer for plastics. Benzotriazoles are UV stable and resistant to biodegradation and oxidation under environmental conditions; as a result they can persist in the environment for a very long time. This kind of substance has the potential to cause damage to aquatic species, compromising their growth and reproduction, due to the disruption of the endocrine system. It can also induce toxic effects in plants such as tomato plants, cucumber seedlings and bush bean plants (Tangtian et al. ; Durjava et al. ). BZ has been found in wastewaters and the subsequently impacted surface waters. This compound is resistant to removal by chlorination, conventional wastewater treatment doi: 10.2166/wst.2014.210

processes, bank ﬁltration, and biologically active ﬁlters. However, it can be degraded by ozone and advanced oxidation processes (Reemtsma et al. ; Mawhinney et al. ). Ultrasound (US) irradiation is a novel advanced oxidation technology that has received considerable attention as a new alternative for removal of persistent organic pollutants (Méndez-Arriaga et al. ; Mahvi ; Chiha et al. ). The propagation of US through a liquid leads to the acoustic cavitation phenomenon. In the cavitation process, bubbles are generated from existing gas nuclei in liquids. They form, grow, and collapse through compression and rarefaction cycles. Temperature and pressure in collapsing bubbles can reach extreme levels (Mahvi ; Chiha et al. ). In aqueous systems, these conditions lead to the thermal dissociation of water vapor into reactive hydroxyl free radicals and hydrogen atoms (Equation (1)); but under different conditions, various other radicals may be present (Chiha et al. ). H2 O þ )))) ! H þ OH

(1)

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Ultrasonic degradation of 1-H-benzotriazole in water

The chemistry involved in the destruction of organic pollutants by US is not the same for all substances and depends on volatility, hydrophobicity, and surface activity of the compound, causing a number of chemical processes. According to the hot spot theory, sonochemical reactions can occur in three different regions: (i) the interior of the collapsing bubbles; (ii) the bubble–liquid interfacial region where the hydroxyl free radical reactions are predominant; and (iii) the bulk of the solution region (Chiha et al. , ). Taking into account this theory, a volatile and hydrophobic pollutant most probably will be decomposed inside the bubble of cavitation, while a non-volatile and hydrophilic compound will be oxidized by the radicals in the interfacial region or bulk solution. Based on the literature and the fact that data on the sonochemical degradation of BZ have not been reported previously, the aim of this study was to evaluate the use of US for BZ degradation, studying the effect of operating parameters such as pH, pollutant initial concentration, applied power, and the presence of co-existing chemical species (dissolved gases: oxygen, nitrogen and ozone; and radical scavengers: alcohols). Additionally, the amount of hydroxyl free radicals generated in the process was determined and its impact on the reaction was analyzed.

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immersed in the solution (probe/horn system). The temperature of the solution was maintained at 25 ± 2 C using a cooling water bath. The amplitude of vibration at the probe tip was controlled at different levels (80, 90, and 100%), corresponding to three nominal powers – 87.50, 101.40, and 114.80 W. Ultrasonic energy dissipated in the reactor was estimated using the calorimetric method; for each system under study the initial temperature rise (T ) was recorded against time (t) in different time intervals. In each sonication run 200 mL of solution were processed, and samples of 1 mL were withdrawn at different time intervals. In the experiments with ozone, samples were quenched using a sodium thiosulfate solution. All the runs were conducted at least three times, and the standard deviations and coefﬁcients of variation of the data were below 5%. W

Analytical methods BZ concentration in water was determined using an Agilent Technologies 1100 HPLC system, with a Phenomenex Luna C-8 column (250 mm × 4.6 mm) and a diode array detector set at 275 nm. The mobile phase was a mixture of methanol and Milli-Q water (30:70, v/v) at a constant ﬂow rate of 1 mL min 1. The column temperature was maintained at 30 C and the injection volume was 100 μL. The OH radical concentration was measured using dimethyl sulfoxide as a capture agent to generate formaldehyde. The formaldehyde was treated with 2,4-dinitrophenylhydrazine to produce formaldehyde-2,4-dinitrophenylhydrazone, which was quantiﬁed using HPLC with an Eclipse Plus C-18 column (150 mm × 4.6 mm) and a mixture of acetonitrile and Milli-Q water (45:55, v/v) as mobile phase. The dissolved organic carbon (DOC) was determined using a Shimadzu TOC-L CPH/CPN instrument. W

MATERIALS AND METHODS Materials All the chemicals used were of analytical grade. Solutions were prepared using Millipore water (18.2 MΩ cm). BZ containing more than 99% of pure compound was purchased from Alfa-Aesar. Ultra-high purity grade oxygen, nitrogen, and air were supplied by Praxair. Ozone was produced from pure oxygen using a laboratory ozone generator (Azco Industries Ltd, VMUS-4S). High performance liquid chromatography (HPLC)-grade methanol and acetonitrile were obtained from Sigma-Aldrich. The pH adjustment was carried out with concentrated solutions of sodium hydroxide and nitric acid obtained from Alfa-Aesar. Ultrasonic reactor Experiments were conducted in a cylindrical glass reactor. An Ultrasonic VCX-500 (Sonics and Materials, USA) was used as US generator. Ultrasonic waves (20 kHz) were emitted from the probe (tip diameter: 13 mm, length: 136 mm, material: titanium alloy) which was completely

RESULTS AND DISCUSSION Effect of initial BZ concentration The effect of BZ initial concentration was studied in the range of 5–20 mg L 1 (41.97–167.88 μmol L 1) and the concentration profiles are shown in Figure 1(a). It can be noticed that the extent of BZ degradation (C/C0) is inversely proportional to the pollutant initial concentration, but the initial degradation rates (depicted in Figure 2), computed as ΔC/Δt over the first minutes of sonication, are higher for higher initial substrate concentrations. A linear relationship was not observed, as expected, for a first-order kinetic law.

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Figure 1

H. Zúñiga-Benítez et al.

Ultrasonic degradation of 1-H-benzotriazole in water

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Effect of different parameters on BZ ultrasonic degradation: (a) effect of BZ initial concentration (frequency: 20 kHz, temperature: 25 ± 2 C, power: 87.5 W, pH: natural (6.5)); (b) W

effect of the applied power (frequency: 20 kHz, temperature: 25 ± 2 C, BZ concentration: 10 mg L 1, pH: natural (6.5)); (c) effect of the initial pH (frequency: 20 kHz, temperature: 25 ± 2 C, BZ concentration: 10 mg L 1, power: 87.5 W); (d) effect of the dissolved gas and (e) effects of different hydroxyl free radical scavengers (alcohols) on the BZ degraW

dation (frequency: 20 kHz, temperature: 25 ± 2 C, BZ initial concentration: 10 mg L 1, power: 87.5 W, pH: 6.5); (f) effect of the dissolved gas on the solution DOC (frequency: W

20 kHz, temperature: 25 þ 2 C, BZ initial concentration: 10 mg L 1, power: 87.5 W, pH: 6.5). W

This observation is consistent with the results that other groups have reported for the sonolytic degradation of other emerging contaminants (Serpone et al. ; Méndez-Arriaga et al. ; Chiha et al. ). It is clear that the sonochemical degradation does not obey a ﬁrst-order kinetics and cannot be characterized by a single rate constant (Chiha et al. ). Table 1 shows some of the physicochemical properties of BZ. From the physical properties, it can be inferred that

this pollutant cannot be degraded by pyrolysis inside the cavitation bubbles (Henry’s law constant is low). In addition, due to its high solubility in water and relatively high octanol/water partition coefﬁcient, BZ preferentially distributes at the interfacial and bulk solution regions. Thus, the OH radicals generated by US are the main species responsible for the pollutant degradation, which most probably occurs not only in the bulk solution but also at the

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Ultrasonic degradation of 1-H-benzotriazole in water

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it is assumed that the rate of adsorption, r1, of pollutant molecules from the bulk solution to the interfacial region of cavitation bubbles is proportional to the pollutant concentration in the bulk solution and (1–θ), where θ corresponds to the occupied ratio of the pollutant molecules in the effective reaction site; and the rate of desorption, r 1, is proportional to θ. Consequently, the rates of r1 and r 1 are expressed by use of the rate constants k1 (min 1) and k 1 (μM min 1) for adsorption and desorption, respectively; as follows (Equations (2) and (3)):

Figure 2

Initial degradation rate as function of the initial concentration of BZ and comparison with predicted values (frequency: 20 kHz, temperature: 25 ± 2 C,

r ¼ k1 C0 (1 θ)

(2)

r 1 ¼ k 1 θ

(3)

Table 1

power: 87.5 W, pH: natural (6.5)).

where C0 is the initial concentration of pollutant in the bulk solution. When the steady state is reached

Physicochemical properties of BZ (SRC PHYSPROP Database, 2013; http://

k1 C0 (1 θ) ¼ k 1 θ

www.srcinc.com/what-we-do/environmental/scientiﬁc-databases.html)

(4)

Therefore, θ can be determined using Equation (5) θ¼

Molecular weight (g gmole 1) 1

119.13 1.98 × 10

Water solubility (mg L ), 25 C Henry’s law constant (atm m mol ), 25 C

1.47 × 10 7

Octanol/water partition coefﬁcient (Log KOW)

1.44

(5)

where K is k1/k 1.When pollutant molecules in the interfacial region of cavitation react with hydroxyl free radicals, the degradation rate, can be expressed by Equation (6) r¼

bubble–liquid interfacial region where hydroxyl free radical concentration is higher (Méndez-Arriaga et al. ). The fact that the degradation rate increases with a higher solute initial concentration shows that the probability of a hydroxyl free radical attack on contaminant molecules increases with increase of substrate in solution, since a higher BZ concentration would imply that initially the hydroxyl attack on contaminant molecule predominates over other reactions such as OH recombination (Chiha et al. , ; Gao et al. ). Additionally, the formation of byproducts could generate additional reactions that could compete with the BZ removal. Models have been developed with the objective of describing the kinetic pattern of sonochemical degradation of non-volatile hydrophilic compounds. This is the case of the heterogeneous kinetic model based on a Langmuirtype mechanism proposed by Okitsu et al. (), in which

KC0 1 þ KC0

kKC 1 þ KC

(6)

where r is the initial degradation rate (μM min 1); k is the pseudo-rate constant (μM min 1); C is the BZ initial concentration (μM) and K is the equilibrium constant (μM 1). Serpone et al. () also proposed a model in which r is the sum of the rates in the bulk solution and the interfacial layer. They considered that the rate of disappearance of the pollutant follows a concentration-independent path and a concentration-dependent course (Chiha et al. ), as shown below (Equation (7)): r ¼ Kb þ

kKC 1 þ KC

(7)

where Kb is a constant representing the rate of decomposition in the bulk liquid (μM min 1), r is the initial degradation rate (μM min 1), k is the pseudo-rate constant

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Ultrasonic degradation of 1-H-benzotriazole in water

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(μM min 1), C is the BZ initial concentration (μM), and K is the equilibrium constant (μM 1). Considering the above, the sonolytic degradation data were analyzed by the non-linear curve fitting analysis method, using Origin 9.0 software. The results of modeling are shown in Table 2 and Figure 2. An excellent fit between the experimental data and both models exists. However, comparing the coefficients of determination (R 2) of each model, it seems that the experimental initial rates for sonochemical degradation of BZ are described better by the Serpone et al. model (R 2 higher). This fact indicates that BZ degradation, at the initial concentration range studied, takes place at both the bubble–liquid interfacial region and in the bulk solution. Figure 3

Hydroxyl free radical content with respect to time of US application, using different input powers (frequency: 20 kHz, temperature: 25 ± 2 C, BZ conW

Effect of the applied ultrasonic power The effect of input power on the BZ degradation is shown in Figure 1(b). The percentages of BZ removal were 42, 46, and 51% under the ultrasonic powers of 87.5, 101.4, and 114.8 W, respectively. Additionally, with increasing power in this range, the initial degradation rate increased, reaching a maximum value around 0.61 μM min 1 for 114.8 W applied power, which meant a 66% increase with respect to the initial rate at the lowest power. It is believed that the number of active cavitation bubbles increases with the increase in the irradiation power. This in turn leads to an increase in the amount of hydroxyl free radical generated (Lim et al. ; Chiha et al. ; Gao et al. ). Experiments carried out to determine the concentration of OH radical based on the methods described by Tai et al. () and Ji et al. (), showed that as the applied power increases, the amount of radical increases, as shown in Figure 3. The increase in the amount of radicals might be responsible for the enhancement in the BZ degradation rates. In addition, a high US

Table 2

Parameters of the models of Okitsu et al. (2005) and Serpone et al. (1994) in the BZ ultrasonic treatment

Model

Parameters

Okitsu et al.

k (μM min 1) ¼ 0.483 16 ± 5.343 03 × 10 4 K (μM) ¼ 0.038 23 ± 1.874 95 × 10 4 R 2 ¼ 0.9970

Serpone et al.

Kb (μM min 1) ¼ 0.015 84 ± 0.008 18 k (μM min 1) ¼ 0.469 16 ± 0.007 18 K (μM) ¼ 0.035 89 ± 0.001 22 R 2 ¼ 0.9999

centration: 10 mg L 1, pH: natural (6.5)) and dissolved gases (power 87.5 W).

power means an increase in the acoustic amplitude, leading to more violent cavitation bubble collapse. Effect of initial pH The effect of solution pH on the sonolytic degradation of BZ was studied by varying the initial pH of the solution. BZ has two pKa values (logarithm of acid dissociation constant), one at 8.2 and the other at 0.4 (Vel Leitner & Roshani ), this means that in aqueous solution BZ can exist in three different conjugate acid–base forms depending on the solution pH. Thus, experiments were carried out at three pH values of 2, 6.5, and 10. In Figure 1(c), the effect of initial solution pH on the BZ sonochemical degradation is depicted. The substrate conversion follows the order; pH 2 > pH 10 > pH 6.5. This behavior can be attributed to the presence of pollutant conjugate acid–base forms in the solution and the extent of dissociation of each case. The ionic fraction of BZ can be calculated using Equation (8) (Chiha et al. ). φions ¼

1 1 þ 10(pKa pH)

(8)

Consequently, at pH values of 2 and 6.5, the compound exists mainly in neutral form, whereas at pH of 10 it is almost completely in anionic form (φions ¼ 0.98), due to deprotonation occurring at the NH group in the ﬁvemembered heterocyclic ring of the BZ molecule. It is expected that the pollutant degradation mainly occurs at the exterior of the cavitation bubbles (interfacial and bulk

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O þ H2 O ! 2 OH

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solution regions). At pH 2 and 6.5, the molecular/neutral form of BZ has a stronger hydrophobic character, and it would distribute mainly in the interface of the cavitation bubbles, where the concentration of hydroxyl free radicals is higher, leading to a faster degradation of the compound. On the other hand, the anionic form of BZ at pH 10 is more reactive toward electrophilic radicals such as OH, enhancing the degradation with respect to the neutral media as has been indicated in the literature (Naik & Moorthy ; Vel Leitner & Roshani ).

These parameters can lead to faster degradation of BZ, mainly due to the greater availability of additional OH radicals. In the case of N2, there is a reduction in the extent of substrate degradation. This can be associated with the fact that dissolved nitrogen present in the aqueous solution can scavenge the free radicals and inhibit the free radical oxidation, as shown in Equations (11)–(16) (Chiha et al. ).

Effects of different dissolved gases

N2 þ OH ! NO2 þ H

(11)

The presence of dissolved gases in the sonochemical treatment of organic substances in water is a parameter that can enhance or inhibit cavitation and may change the reaction efficiency, promoting radical species not only from sonolysis of the solvent but also from the dissolved gases involved (Méndez-Arriaga et al. ; Chiha et al. ). In order to study the influence of the dissolved gases on the degradation of BZ, experiments with ozone, oxygen, and nitrogen were carried out during the US irradiation. In the case of O2 and N2, the flow rate was 762 mL min 1, while in the experiments with ozone, pure oxygen (762 mL min 1) was used to generate 3.5 g h 1 of O3. Figure 1(d) shows the time dependence of BZ degradation under O3, O2, and N2 atmospheres. It can be seen from the figure that the order of degradation ratio is US þ O3 >O3 alone >US þ O2 > US alone >US þ N2. In the presence of ozone, the coupling of ultrasonic irradiation with ozonation slightly improves the pollutant degradation compared to ozonation alone, which appears to be the main oxidizing agent in this case. In both cases the same dose of O3 was employed, but combination of US and ozone leads to generation of more hydroxyl free radicals in comparison with the other gases as shown in Figure 3. In sonolysis, ozone is decomposed thermolytically in the vapor phase of the cavitation bubbles, generating additional hydroxyl radicals (Weavers et al. ). Although in terms of C/C0 a significant difference between USþ O3 and US alone is not evident, in terms of mineralization, as Figure 1(f) shows, the presence of US accelerates the DOC reduction. On the other hand, under an oxygen atmosphere extra hydroxyl free radicals are generated by molecular oxygen dissociation inside the cavitation bubble (Equations (9) and (10)) (Pétrier et al. ).

N2 þ O ! NO þ N

(12)

O2 þ N ! NO þ O

(13)

NO þ O ! NO2

(14)

NO þ OH ! HNO2

(15)

NO2 þ OH ! HNO3

(16)

O2 ! 2O

(9)

(10)

Figure 3 also veriﬁes that oxygen promotes the radical generation, while nitrogen reduces considerably the amount of OH radicals.

Effects of hydroxyl radical scavengers The effect of the presence of four alcohols (methanol, ethanol, isopropyl alcohol, and tert-butyl alcohol) in the sonolytic degradation of BZ was investigated. These alcohols, as well as other organic species of low molecular weight can act as quenchers in reactions with OH radicals (Behnajady et al. ). Figure 1(e) shows that all of them inhibit the degradation of BZ, which is evidence of the role of the hydroxyl radicals as the main contaminant oxidant. The inhibition of the 1-H degradation is due to •OH competitive reactions with BZ and alcohols. The scavenging activity follows the order tert-butyl alcohol > isopropyl alcohol > ethanol > methanol, which can be attributed to the possible reaction of hydroxyl free radicals with hydrogens bonded to carbon in alcohol molecules (Behnajady et al. ). There are nine, seven, ﬁve, and three hydrogen atoms, in the tert-butyl alcohol, isopropyl alcohol, ethanol,

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and methanol molecules, respectively. On the other hand, this scavenging activity order was in agreement with the order of alcohol specific heat capacities, which could indicate that an alcohol with larger specific heat exhibits a greater inhibition effect. This may be attributed to the fact that large specific heat capacities would reduce the maximum size of micro-bubbles and the interfacial temperatures achieved from bubble collapse, which could produce fewer OH radicals (Gao et al. ). Mineralization of BZ The change in the DOC content of the solution under the different conditions evaluated is shown in Figure 1(f). It can be noted that there was not a significant reduction in this parameter. In most of the experiments, around 10% reduction in DOC was achieved. This indicates that BZ is being transformed into organic byproducts with low volatility and high hydrophilicity, and thermal decomposition inside the bubbles was insignificant. It has been reported that reactions of hydroxyl free radicals with BZ could give hydroxycyclohexadienyl radicals, by addition of OH radical in the benzene ring present in the BZ molecule (Naik & Moorthy ). Only in the presence of ozone did the percent of DOC reduction reach around 76% in just 20 min; it could indicate that organic solutes were mineralized due to the combined oxidation with ozone and US.

CONCLUSIONS Removal of the endocrine-disrupting BZ by 20 kHz ultrasonic irradiation was studied. Hydroxyl radicals generated by collapse of cavitation bubbles during the sonochemical treatment are responsible for the BZ degradation. Kinetic equations based on a Langmuir-type mechanism were used to model the sonochemical degradation of the BZ. The expression developed by Serpone et al. () showed a better ﬁt with the experimental data; in comparison with the Okitsu et al. () model, indicating that pollutant degradation takes place at both the bubble–liquid interfacial region and in the bulk solution. The degradation rate increased with increasing ultrasonic power from 87.5 to 114.8 W due to increase in the amount of OH. Acidic media was the most favorable condition for the degradation of BZ. In addition, in the presence of O3 and O2, the degradation rate was higher since they promote the hydroxyl radical generation whereas N2 acted as an OH scavenger and inhibited the BZ removal. It is very difﬁcult to

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mineralize BZ by ultrasonic treatment alone; however, in the presence of ozone a reduction of around 76% of the total organic carbon in the solution was achieved.

ACKNOWLEDGEMENTS The authors wish to thank the Canadian Bureau for International Education (CBIE) and its ELAP program; the Colombian Administrative Department of Science, Technology and Innovation (COLCIENCIAS); the University of Saskatchewan and the sustainability fund 2011–2012 from the Vice-Rectory for Research at the University of Antioquia for the support of this work.

REFERENCES Behnajady, M. A., Modirshahla, N., Shokri, M. & Vadhi, B.  Effect of operational parameters on degradation of Malachite Green by ultrasonic irradiation. Ultrasonics Sonochemistry 15 (6), 1009–1014. Bolong, N., Ismail, A. F., Salim, M. R. & Matsuura, T.  A review of the effects of emerging contaminants in wastewater and options for their removal. Desalination 239 (1–3), 229–246. Chiha, M., Merouani, S., Hamdaoui, O., Baup, S., Gondrexon, N. & Pétrier, C.  Modeling of ultrasonic degradation of nonvolatile organic compounds by Langmuir-type kinetics. Ultrasonics Sonochemistry 17 (5), 773–782. Chiha, M., Hamdaoui, O., Baup, S. & Gondrexon, N.  Sonolytic degradation of endocrine disrupting chemical 4-cumylphenol in water. Ultrasonics Sonochemistry 18 (5), 943–950. Durjava, M. K., Kolar, B., Arnus, L., Papa, E., Kovarich, S., Sahlin, U. & Peijnenburg, W.  Experimental assessment of the environmental fate and effects of triazoles and benzotriazole. Alternatives to Laboratory Animals: ATLA 41 (1), 65–75. Gao, Y., Gao, N., Deng, Y., Gu, J., Gu, Y. & Zhang, D.  Factor affecting sonolytic degradation of sulfamethazine in water. Ultrasonics Sonochemistry 20 (6), 1401–1407. Ji, G., Zhang, B. & Wu, Y.  Combined ultrasound/ozone degradation of carbazole in APG1214 surfactant solution. Journal of Hazardous Materials 225–226, 1–7. Lim, M. H., Kim, S. H., Kim, Y. U. & Khim, J.  Sonolysis of chlorinated compounds in aqueous solution. Ultrasonics Sonochemistry 14 (2), 93–98. Mahvi, A. H.  Application of ultrasonic technology for water and wastewater treatment. Iranian Journal of Public Health 38 (2), 1–17. Mawhinney, D. B., Vanderford, B. J. & Snyder, S. h. A.  Transformation of 1H-benzotriazole by ozone in aqueous solution. Environmental Science and Technology 46 (13), 7102–7111. Méndez-Arriaga, F., Torres-Palma, R., Pétrier, C., Esplugas, S., Gimenez, J. & Pulgarín, C.  Ultrasonic treatment of

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water contaminated with ibuprofen. Water Research 42 (16), 4243–4248. Naik, D. B. & Moorthy, P. N.  Studies on the transient species formed in the pulse radiolysis of benzotriazole. Radiation Physics and Chemistry 46 (3), 353–357. Okitsu, K., Iwasaki, K., Yobiko, Y., Bandow, H., Nishimura, R. & Maeda, Y.  Sonochemical degradation of azo dyes in aqueous solution: a new heterogeneous kinetics model taking into account the local concentration of OH radicals and azo dyes. Ultrasonics Sonochemistry 12 (4), 255–262. Pétrier, C., Combet, E. & Mason, T.  Oxygen-induced concurrent ultrasonic degradation of volatile and non-volatile aromatic compounds. Ultrasonic Sonochemistry 14 (2), 117–121. Reemtsma, T. h., Miehe, U., Duennbier, U. & Jekel, M.  Polar pollutants in municipal wastewater and the water cycle: occurrence and removal of benzotriazoles. Water Research 44 (2), 596–604. Serpone, N., Terzian, R., Hidaka, H. & Pelizetti, E.  Ultrasonic induced dehalogenation and oxidation of 2-, 3and 4-CP in air-equilibrated aqueous media, similarities with

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irradiated semiconductor particles. Journal of Physical Chemistry 98, 2634–2640. Tai, C., Peng, J., Liu, J., Jiang, G. & Zou, H.  Determination of hydroxyl radicals in advanced oxidation processes with dimethyl sulfoxide trapping and liquid chromatography. Analytica Chimica Acta 527 (1), 73–80. Tangtian, H., Bo, L., Wenhua, L., Shin, P. K. & Wu, R. S.  Estrogenic potential of benzotriazole on marine medaka (Oryzias melastigma). Ecotoxicology and Environmental Safety 80, 327–332. Vel Leitner, N. K. & Roshani, B.  Kinetic of benzotriazole oxidation by ozone and hydroxyl radical. Water Research 44 (6), 2058–2066. Weavers, L. K., Ling, F. H. & Hoffmann, M. R.  Aromatic compound degradation in water using a combination of sonolysis and ozonolysis. Environmental Science Technology 32 (18), 2727–2733. Xu, J., Li, L., Guo, C., Zhang, Y. & Wang, S.  Removal of Benzotriazole from solution by BiOBr photocatalysis under simulated solar irradiation. Chemical Engineering Journal 221, 230–237.

First received 20 December 2013; accepted in revised form 16 April 2014. Available online 30 April 2014

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New equation for the computation of flow velocity in partially filled pipes arranged in parallel Lotfi Zeghadnia, Lakhdar Djemili, Larbi Houichi and Nouredin Rezgui

ABSTRACT This paper presents a new approach for the computation of flow velocity in pipes arranged in parallel based on an analytic development. The estimation of the flow parameters using existing methods requires trial and error procedures. The assessment of flow velocity is of great importance in flow measurement methods and in the design of drainage networks, among others. In drainage network design, the flow is mostly of free surface type. A new method is developed to eliminate the need for trial methods, where the computation of the flow velocity becomes easy, simple, and direct with zero deviation compared to Manning equation results and other approaches such as that have been considered as the best existing solutions. This research work shows that these approaches lack accuracy and do not cover the entire range of flow surface angles: 0 θ 360 . W

Key words

Lotfi Zeghadnia (corresponding author) Nouredin Rezgui Department of Civil Engineering, Faculty of Sciences & Technology, University of Mohamed Cherif Messaadia, Souk Ahras, Algeria E-mail: Zeghadnia_lotfi@yahoo.fr Lotfi Zeghadnia Lakhdar Djemili Department of Hydraulics, Faculty of Engineering Science, University of Badji-Mokhtar, Annaba, Algeria

| circular pipe, flow velocity, free surface flow, Manning equation, uniform and steady flow

Larbi Houichi Department of Hydraulics, Faculty of Engineering Science, University of Elhadj Lakhdar, Batna, Algeria

INTRODUCTION Flow velocity calculation is an important task for hydraulic engineers, especially in the design of open channel conveyance for irrigation, drainage, and sewer networks. Flow in the latter is usually under free surface conditions, in order to avoid a decrease of the water cross-sectional area and thereby the decrease of the flow efficiency. The Manning model is considered as the best model to describe free surface flow. Several authors, such as Chow (), Henderson (), Metcalf & Eddy Inc. (), Carlier (), Swamee (), Swamee & Rathie (), and Hager (), have discussed the model at length. We considered in this study pipes with a circular crosssection as these are the preferred form for sewer system design. Subwatersheds can be arranged in series or in parallel. Similarly, pipes in a sewer or drainage system can be arranged in series or in parallel. Using Manning’s equation, the computation of the flow velocity and water surface angle is not direct and needs to go through trial methods with heavy computations. Based on the Manning model, several researchers have been attempting to propose an explicit solution for free surface flow computation. Among these are Saatçi (), Giroud et al. (), and Akgiray (, ) who have tried to eliminate the need for iterative and trial methods for water surface angle ranging between 0 and 302.41 . W

doi: 10.2166/wst.2014.199

Transition into a pressurised flow regime may occur during intense rain events as inflow exceeds the transport capacity of the system in free flow mode. A number of authors considered that it is reliable to simulate flow in a pressurised pipe as surface flow using the Preissman slot pipe method to study the transition from free surface flow to a surcharged state in the pipe (Cunge et al. ; Garcia Navarro et al. ; Capart et al. ; Ji ; Trajkovic et al. ; Ferreri et al. ). Other approaches are exemplified by a number of authors. They may be classified as rigid column method (Wiggert ; Hamam & McCorquodale ; Li & McCorquodale ; Zhou et al. ; Vasconcelos & Wright ), or full-dynamic models (Song et al. ; Cardle & Song ; Guo & Song ). Others preferred experimental methods, such as Jose & Steven (), and Ciraolo & Ferreri (), preferred experimental methods. In this study, based on the Manning model and without taking into consideration the rapid filling of the pipe, a new approach is being proposed. It is much simpler and more accurate than the other methods for the computation of the flow velocity in partially filled pipes arranged in parallel form, for all the range of surface water angle: 0 θ 360 . W

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Water Science & Technology

THEORY

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V¼

1 D 2=3 ðθ sin θÞ 2=3 1=2 s n 4 θ

(4)

A¼

D2 ðθ sinðθÞÞ 8

(5)

Manning equation For a long time, the Manning equation (Manning ), has been frequently considered as the best formula to compute free surface flow owing to its simplicity. The Manning equation can be used for uniform flow. Graphs and tables are established to facilitate the application of the Manning equation for the estimation of flow characteristics (Camp ; Swarna & Modak ; McGhee ). Manning’s equation can be written as follows: 1 2=3 Q ¼ Rh AS1=2 n V¼

(1)

1 2=3 1=2 R S n h

(2) 1=2

where Q is the flow rate (m3/s), Rh is the hydraulic radius (m), n is the pipe roughness coefficient, or Manning n (s/m1/3), A is the cross-sectional flow area (m2), S is the slope of pipe bottom, dimensionless, and V is the flow velocity (m/s). To use the Manning model some hypotheses must be assumed: the flow must be steady and uniform, where the slope, cross-sectional flow area, velocity are not related to time, and constant throughout the length of pipe being analysed (Carlier ). Equations (1) and (2) can be rewritten as a function of the water surface angle of the pipe, as shown in Figure 1, as follows: #1=3 1=3 " 1 D8 ðθ sinθÞ5 Q¼ s1=2 n 213 θ2

(3)

P¼θ

D 2

(6)

A D sinðθÞ 1 Rh ¼ ¼ P 4 θ

(7)

where D is the pipe diameter (m), P is the wetted perimeter (m), and θ is the water surface angle (radian). According to the equations mentioned above, the computation of the ﬂow velocity is not direct, and needs to go through iterations. Several authors have attempted to propose approximate approaches to eliminate the need for trial procedures. Among these, we note those of Saatçi (), Giroud et al. (), and Akgiray (). Saatçi’s approach is based on two steps. In the ﬁrst, we should estimate the water surface angle using the following formula:

θSaatçi

vffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi sffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi u rffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi u 3π t πQn 1 1 ¼ 2 D8=3 S0:5

(8)

In the second step, Equation (4) is used to calculate the flow velocity. The equations of Saatçi cannot be used for all the range of θ. They can be used only for the range: θ 265 (Saatçi ). Giroud et al. () and Akgiray () have also proposed approaches in order to improve Saatçi’s equation. The Giroud model proposes a direct relationship between average flow velocity and the flow rate for θ between 0 and 301.41 . This approach produces an estimate with a maximum deviation of less than 3%. As well, Akgiray tried to improve the approaches presented earlier. He proposed approximate solutions from Manning’s equation for four types of problems: W

1. 2. 3. 4.

Figure 1

Water surface angle.

Given Given Given Given

Q, D, and S, find h/D and/or V. Q, D, and V, find h/D and/or S. V, D, and S, find h/D and/or Q. Q, V, and S, find h/D and/or D.

The four problems mentioned above are studied for two cases, when the Manning coefﬁcient n is constant, and when n varies with the ﬂow depth, as documented by Camp

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Computation of ﬂow velocity in partially ﬁlled pipes

Water Science & Technology

(). In the latter, both variables are considered to be known. In this research, we are interested in the ﬁrst problem, for which Akgiray () proposed the following equations to assess the water surface angle: θ ¼ 2 × 65=13 K3=13

0:8 1 0:946 1 þ sin ð2:98KÞ 2K

(9)

where K¼

Qn D8=3 S0:5

(10)

Equation (9) is proposed for θ between 0 and 301.41 with maximum error equal to 0.72%. The ﬂow velocity can be estimated according to Equations (9) and (4), where the maximum deviation is 0.29%. W

•

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The pipe N1-N2 collects water from the equivalent watershed CC3 (which takes on number 3). CC3 ¼ CC2 þ CC1

(11)

The ﬂow Q can be evaluated using current methods, such as the rational method or runoff curve number method (Viessman & Lewis ). Following are the four possible types of problems to compute the ﬂow velocity in relation to the reference pipe: 1. The computation of V3 as a function of the characteristics (pipe 01 is the reference pipe). 2. The computation of V3 as a function of the characteristics (pipe 02 is the reference pipe). 3. The computation of V2 as a function of the characteristics (pipe 01 is the reference pipe). 4. The computation of V1 as a function of the characteristics (pipe 02 is the reference pipe).

pipe 01 pipe 02 pipe 01 pipe 02

THE NEW APPROACH

Cases 1 and 2

Flow velocity

Flow in pipes is steady and uniform, which means that the ﬂow characteristics are constant in time and in space (throughout the length of pipe being analysed). Let us consider that the pipe M2-N1 (or pipe 02) is the reference pipe, with known parameters. As such, the diameter D2, hydraulic radius Rh2, surface water angle θ2, crosssectional water A2 and slope S2, are known data. The slope S3 and roughness n3 are considered as known parameters for the pipe N1-N2 (or pipe 03). Equation (4) can also be written as the following equation (Zeghadnia et al. ):

Subwatershed arranged in parallel Subwatersheds may be arranged in series or in parallel. This study focuses on the second case, where the pipes are arranged in parallel as shown in Figure 2.

• •

The pipe M1-N1 collects water from subwatershed CC1 (which takes on number 1). The pipe M2-N1 collects water from the equivalent watershed CC2 (which takes on number 2).

V¼

1=2 3 2 !1=5 S 2Q θ 2=5 D n

(12)

where θ is the water surface angle in radians as shown in Figure 2. Q1 is transported in pipe M1-N1 and produced in subwatershed CC1, Q2 is transported in pipe M2-N1 and produced in subwatershed CC2, and Q3 is transported in pipe N1-N2 and produced in subwatershed CC3.

Figure 2

Subwatershed and pipes arranged in parallel.

Q3 > Q1

(13)

Q3 > Q2

(14)

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Computation of ﬂow velocity in partially ﬁlled pipes

The ratios Rq32 between Q3 and Q2 and Rq31 between Q3 and Q1, are given by the following equations: Q3 ¼ Rq32 Q2

(15)

Q3 ¼ Rq31 Q1

(16)

where Q3 ¼ A3 V3

(17)

Q1 ¼ A1 V1

(18)

Water Science & Technology

)a ¼

V33

n3 s0:5 3

4 D2

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2 (26)

The combination of Equations (24) and (26) gives the following equation: 1=4

V3 ¼ ðrQ32 V2 Þ

S0:5 3 n3

3=4

D2 4

1=2 (27)

Equation (27) allows the computation of the ﬂow velocity in just full pipe (pipe N1-N2) as a function of the reference pipe characteristics. In the case of partially ﬁlled pipe and according to Equation (12), the previous formula can be rewritten as follows:

From the inequalities shown in Equations (13) and (14) for just full pipe (under atmospheric pressure), we obtain the following equation:

!3=20 1=4 0:5 3=4 S3 Q3 n2 2Q2 2=5 V3 ¼ Q2 n3 D2 θ2 S0:5 2

A3 ¼ bA1

Similarly, in the case where pipe M1-N1 is the reference pipe, the ﬂow velocity can be calculated as follows:

(19)

This gives: D23 ¼ bD21

(20)

Similarly, A3 ¼ aA2

(21)

where

!3=20 1=4 0:5 3=4 S3 Q3 n1 2Q1 2=5 V3 ¼ Q1 n3 D1 θ1 S0:5 1

(28)

(29)

Equations (28) and (29) give the exact value of ﬂow velocity in pipe N1-N2 as a function of the reference pipe parameters, both for the ﬁrst and for the second case, where the maximum deviation is zero as compared to Equation (4) results. Cases 3 and 4

D23

aD22

(22)

Based on Equations (15), (17), and (21), the ratio Rq32 becomes as follows: Rq32 ¼

aV3 V2

(23)

In these cases, the reference pipe characteristics are known. Here, we will write the characteristics of the ﬁrst pipe using the characteristics of the second pipe (reference pipe) as will be shown in the following sections. Using Equations (20) and (22), we obtain the following equations: aD22 ¼ bD21

(30)

which yields: Rq32 V2 V3 ¼ a

(24)

2=3 s0:5 D2 3 a1=3 n3 4

a 0:5 D2 b

(31)

Using Equation (31), the velocity equation can be written as follows:

Using Equation (22), (2) becomes as follows:

V3 ¼

D1 ¼

(25)

V1 ¼

2=3 1=3 s0:5 D2 a 1 b n1 4

(32)

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Water Science & Technology

Based on Equations (16), (17), and (19), the ratio Rq31 becomes as follows: Rq31 ¼

bV3 V1

a V2 Q1 ¼ b V1 Q2

(33)

θ1 and θ2

2014

(35)

where

Rq31 V1 b

Table 1

70.1

Also, using Equations (24) and (34), we can get the following equation:

which yields: V3 ¼

Rq12 ¼

(34)

Q1 Q2

(36)

Accuracy test of Equation (38) compared to the recent approaches and Equation (4)

Manning equation (4)

Proposed equation (38)

Error %

Saatçi equation error %

Giroud equation error %

Akgiray equation error %

0.002175

89.34740

0.004

3.82208 × 10 3

0.005480

85.94367

0.004

1.83529 × 10 3

0.009410

9.896 × 10 6

83.46836

0.005

1.30634 × 10 3

0.013808

81.45177

0.007

3.97250 × 10 3

0.018592

79.71937

0.009

5.74054 × 10 3

0.023705

78.18411

0.012

8.23439 × 10 3

0.029111

6.398 × 10 6

76.79537

0.015

1.16579 × 10 2

0.034779

75.52073

0.018

1.50708 × 10 2

0.040686

74.33798

0.022

1.90265 × 10 2

0.046813

73.23115

0.027

2.34751 × 10 2

0.117604

64.62856

0.096

9.48772 × 10 2

0.245945

55.56500

0.276

0.295587

0.341063

50.64856

0.437

0.493892

0.807894

32.40590

1.190

2.001798

0.911895

28.60162

1.272

2.455752

1.111704

20.80760

1.244

3.462290

1.250781

14.60113

1.027

4.285882

1.336877

10.19018

0.791

4.857391

1.641918

14.59881

0.836

7.359813

1.689098

21.95952

1.192

7.810522

100 120 135 145 190 200

1.792165

6.651 × 10

1.804408

1.809674

6.587 × 10 6

119.68460

1.220

8.683651

1.778588

Not applicable

1.468

8.126489

1.758345

Not applicable

2.694

8.254137

1.739430

Not applicable

3.803

9.677319

1.663581

Not applicable

8.209

17.299310

1.633169

Not applicable

10.013

20.004720

1.588002

Not applicable

12.768

23.802130

235

245 256

290 300 308 335 345 360

60.96910

1.749

8.712387

80.98525

1.597

8.748280

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Computation of ﬂow velocity in partially ﬁlled pipes

Using Equations (32), (35), and (36), we may estimate the velocity using the characteristics of the second pipe for just full pipe as follows: V1 ¼ ðRq12 V2 Þ1=4

S0:5 1 n1

3=4

D2 4

1=2 (37)

For the case of partially full pipe, and according to Equation (12), Equation (37) becomes: V1 ¼

Q1 Q2

1=4

S0:5 1 n1

3=4

n2 S0:5 2

!3=20

2Q2 D2 θ 2

2=5 (38)

Equation (38) is the best formula as compared to existing methods with regard to Equation (4). This equation produces a maximum error of zero as shown in Table 1. For the case when pipe 01 is the reference pipe, the same result can be obtained as follows: V2 ¼

Q2 Q1

1=4

S0:5 2 n2

3=4

n1 S0:5 1

!3=20

2Q1 D1 θ 1

2=5 (39)

Accuracy test Table 1 shows that Equation (38) is an excellent formula, with a maximum deviation of 9:89 × 10 6 % ≅ 0, as compared to the Manning equation (Equation (4)) results. Existing methods, such as those of Saatçi (), Giroud et al. (), and Akgiray (), produce large error with limited range of application. For example, maximum deviation ΔV/V for Saatçi’s equation is much higher and is not applicable for the entire range of θ. As well, the Giroud equation results for θ between 0 and 360 produce a maximum error of 12.76%. For the Akgiray formula, the maximum deviation is 23.80%. On the other hand, the proposed approach in this study using Equation (38) for the ﬁrst case or Equation (39) for the second case is much more accurate, with a maximum error of zero. So far, it is the best existing formula as compared to Equation (4). W

CONCLUSIONS This paper presents a novel approach for the explicit computation of the ﬂow velocity in partially ﬁlled circular-sections for four possible types of problems using known characteristics of a reference pipe. The proposed solution may be

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used for full circular sections. Its results are much better than those of existing models, where the deviation between the proposed equation and Manning results is zero as shown in Table 1. The proposed equation should be of great importance and application for concerned engineers and professionals.

ACKNOWLEDGEMENT The authors would like to thank Prof. Djebbar Yassine for his help and his positive collaboration.

REFERENCES Akgiray, Ö.  Simple formula for velocity, depth of flow and slope calculations in partially filled circular pipes. Environ. Eng. Sci. 21 (3), 371–385. Akgiray, Ö.  Explicit solutions of the Manning equation in partially filled circular pipes. Environ. Eng. Sci. 32, 490–499. Camp, T. R.  Design of sewers to facilitate flow. J. Sewage Works 18 (1), 3–16. Capart, H., Sillen, X. & Zech, Y.  Numerical and experimental water transients in sewer pipes. J. Hydraul. Res. 35 (5), 659–670. Cardle, J. A. & Song, C. S. S.  Mathematical modeling of unsteady flow in storm sewers. Int. J. Eng. Fluid Mech. 1 (4), 495–518. Carlier, M.  Hydraulique Générale et Appliquée. Eyrolles, Paris. Chow, V.-T.  Open Channel Hydraulics. McGraw-Hill, New York. Ciraolo, G. & Ferreri, G. B.  Experimental investigation on pressurization transient of a drainage sewer. In: Proceedings of 5th International Symposium on Environmental Hydraulics – ISEH-V, Tempe, Arizona (USA), December, on CD-ROM. Cunge, J. A., Jr., Holly, F. M. & Verwey, A.  Practical Aspects of Computational River Hydraulics. Pitman, London. Ferreri, G. B., Freni, G. & Tomaselli, P.  Ability of Preissmann slot scheme to simulate smooth pressurisation transient in sewers. Water Sci. Technol. 62 (8), 1848–1858. Garcia-Navarro, P., Priestley, A. & Alcrudo, F.  Implicit method for water flow modeling in channels and pipes. J. Hydraul. Res. 32 (5), 721–742. Giroud, J. P., Palmer, B. & Dove, J. E.  Calculation of flow velocity in pipes as function of flow rate. J. Geosynth. Int. 7 (4–6), 583–600. Guo, Q. & Song, C. S. S.  Surging in urban storm drainage systems. J. Hydraul. Eng. 116 (12), 1523–1537. Hager, W. H.  Wastewater Hydraulics Theory and Practice, 2nd edn. Springer, London. Hamam, M. A. & McCorquodale, J. A.  Transient conditions in the transition from gravity to surcharged sewer flow. Can. J. Civ. Eng. 9, 189–196. Henderson, F. M.  Open Channel Flow. Macmillan, NewYork.

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Ji, Z.  General hydrodynamic model for sewer/channel network systems. J. Hydraul. Eng. 124 (3), 307–315. Jose, G. V. & Steven, J. W.  Experimental investigation of surges in a stormwater storage tunnel. J. Hydraul. Eng. 131 (10), 853–861. Li, J. & McCorquodale, A.  Modeling mixed flow in storm sewers. J. Hydraul. Eng. 125 (11), 1170–1180. Manning, R.  On the flow of water in open channels and pipes. Trans. Inst. Civ. Eng. Ireland 20, 161–207. McGhee, T. J.  Water Supply and Sewerage, 6th edn. McGrawHill, New York. Metcalf & Eddy Inc.  Wastewater Engineering: Collection and Pumping of Wastewater. McGraw-Hill, New York. Saatçi, A.  Velocity and depth of flow calculations in partially filled pipes. J. Environ. Eng. 116 (6), 1202–1208. Song, C. S. S., Cardle, J. A. & Leung, K. S.  Transient mixed flow models for storm sewers. J. Hydraul. Eng. 109 (11), 1487–1504. Swamee, P. K.  Normal depth equations for irrigation canals. ASCE J. Hydraul. Div. 102 (5), 657–664. Swamee, P. K. & Rathie, P. N.  Exact solution for normal depth problem. J. Hydraul. Res. 42 (5), 657–664.

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Swarna, V. & Modak, P.  Graphs for hydraulic design of sanitary sewers. J. Environ. Eng. 116 (3), 561–574. Trajkovic, B., Ivetic, M., Calomino, F. & D’Ippolito, A.  Investigation of transition from free surface to pressurized flow in a circular pipe. Water Sci. Technol. 39 (9), 105–112. Vasconcelos, J. G. & Wright, S. J.  Surges associated with expulsion in near-horizontal pipelines. In: Proc. FEDSM03. 4th ASME-JSME Joint Fluids Engineering Conf., Honolulu, USA. Viessman, W. & Lewis, G. L.  Introduction to Hydrology, 5th edn. Prentice-Hall, Upper Saddle River, NJ, USA. Wiggert, D. C.  Transient flow in free-surface, pressurized systems. J. Hydraul. Div. Am. Soc. Civ. Eng. 98 (1), 11–27. Zeghadnia, L., Djemili, L., Houichi, L. & Rezgui, N.  Détermination de la vitesse et la hauteur normale dans une conduite partiellement remplie (Estimation of the flow velocity and normal depth in partially filled pipe). J. EJSR. 37 (4), 561–566. Zhou, F., Hicks, F. E. & Steffler, P. M.  Transient flow in a rapidly filling horizontal pipe containing trapped air. J. Hydraul. Eng. 128 (6), 625–634.

First received 24 July 2013; accepted in revised form 10 April 2014. Available online 30 April 2014

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The regularity of wind-induced sediment resuspension in Meiliang Bay of Lake Taihu Jianjian Wang, Yong Pang, Yiping Li, Yawen Huang, Junjie Jia, Peng Zhang and Xiaopeng Kou

ABSTRACT Contaminants released by wind-induced sediment resuspension could inﬂuence the water quality in shallow lakes. This study aims to reveal the quantitative relationship between wind speed (v) and sediment resuspension rate (r) in Meiliang Bay of Lake Taihu. The study was conducted in three steps. First, the in situ wind speed and current velocity were measured over a period of 2 days in Meiliang Bay to establish the relationship between wind and hydrodynamic conditions; second, an indoor experiment was conducted in a cylindrical simulator with sediment from the study area to determine sediment resuspension rates under different hydrodynamic conditions; and third, linkages between sediment resuspension and wind were determined. The average sediment resuspension rate was highly correlated with the wind speed (R 2 ¼ 0.99), and was expressed by r ¼ 20.72v 2.034 at wind speeds in the range of 0–14 m/s. The critical wind speed for sediment resuspension is about 7 m/s. Under these conditions, the average resuspension rate could reach 1,000 g/(m2 d), with a total phosphorus release rate of 1.1 g/(m2 d) and a total nitrogen release rate of 18.1 g/(m2 d). Key words

Jianjian Wang (corresponding author) Yong Pang Yiping Li Yawen Huang Junjie Jia Peng Zhang Xiaopeng Kou College of Environment, Hohai University, 1 Xikang Road, Nanjing 210098, China E-mail: hhuwangjianjian@126.com; wangnpy@163.com Yong Pang Yiping Li Key Laboratory for Integrated Regulation and Resources Exploitation on Shallow Lakes, Ministry of Education, Hohai University, Nanjing 210098, China

| cylindrical simulator, sediment resuspension, shear stress, wind

INTRODUCTION Lake eutrophication is a serious environmental problem in China, especially in shallow lakes. Lake Taihu is highly eutrophic, despite many restoration efforts, including the use of macrophytes, sediment dredging, wetland construction, water transfer, and increased nutrient recycling (Xiao et al. ; Ye et al. ; Zhai et al. ). The lake’s water quality has not shown any significant improvement, and Meiliang Bay in the northwestern portion of the lake remains especially impaired (Qin ; Zhao et al. ). Frequent resuspension of sediments is recognized as an important process which impedes the recovery of eutrophic lakes, especially when external nutrient sources are controlled (Ahlgren et al. ; Christian ). Sediment resuspension occurs when bottom shear stress is sufficiently large to disrupt the cohesion of benthic sediment. Wind has a significant effect on the flow velocity in shallow lakes, such as Lake Taihu (Maxam & Webber ). Wind-induced currents frequently disturb the water–sediment interface, and often cause significant increases of suspended sediment concentration (SSC) in overlying water. The effect of strong winds on SSC in shallow lakes has been well documented doi: 10.2166/wst.2014.217

(e.g., Luettich et al. ; Bengtsson & Hellström ; Arfi et al. ; Qin et al. ; Zhu et al. ). Field observation of sediment resuspension has been achieved under different wind conditions (James et al. ; Horppila & Nurminen ; Hu et al. ). The Y-shape apparatus, the annular tank, and the particle entrainment simulator were used to produce various degrees of water turbulence in indoor experiments (Cantwell & Burgess ; Li et al. ; Li ; Wan et al. ; Huai et al. ), but there are still some disadvantages in these methods. Field observation is hard to duplicate under the same environment conditions, and compared to the natural process of sediment resuspension in shallow lakes, the flow velocity created by these simulative devices is more uniform, and the natural characteristics of the sediment may be destroyed. Models have also proved useful for predicting the effect of suspended sediment (Hamilton & Mitchell, ; Bailey & Hamilton ). In this study, field observations were made and an improved indoor simulator was used. During the field observation, an acoustic Doppler velocimeter (ADV, SonTek,

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USA) was used to measure three-dimensional ﬂow velocity, while a portable aerovane was used to measure wind speed at the same time. In order to understand the relationship between current velocity and sediment resuspension rate in this area, an indoor dynamic simulation experiment was conducted in a self-designed cylindrical simulator (Chinese Invention Patent, application number: 201310236315.1) with sediment taken from the experimental area. The cylindrical simulator was especially designed to simulate the ﬂow structure observed in Lake Taihu. Meanwhile, a sediment sampler (Chinese Invention Patent, application number: 201310184563.6) was designed in this study to preserve the vertical structure of sediment while sampling. The relationship between wind speed, current velocity and sediment resuspension was developed and validated to provide a theoretical basis for understanding the process of sediment resuspension and estimating the amount of sediment contaminants released in Lake Taihu.

MATERIALS AND METHODS Experimental design

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Table 1

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between wind and sediment resuspension through the intermediate variable of shear stress. This study had two major parts: field observations and an indoor experiment. The design was as follows. (1) An ADV and a portable aerovane were used to observe the lake flow and wind in the field synchronously, the shear stress was calculated, and then the relationship between the wind and the shear stress was established. (2) Sediment resuspension status was simulated under different shear stresses in the laboratory to establish the relationship between the shear stress and sediment resuspension. (3) The laboratory results were translated to the field through shear stress to determine the in situ relationship between wind and sediment resuspension.

Field observation Lake Taihu is a large shallow lake with a surface area of 2,338 km2 and an average depth of 1.9 m. Table 1 shows the annual average hydrological and meteorological characteristics (Qin et al. ; Huang & Xu ). For Lake Taihu, the average annual water temperature is 17.3 C with a range of 4.8–29.2 C. The vertical temperature drop mainly varies from 0 to 1 C (Zhao et al. ). A northwest wind prevails in the winter while a southeast wind prevails in the summer. But in spring and autumn, the wind direction is changeable. Generally, the wind speed varies between 0 and 10 m/s, with an average speed of 4.3 m/s in spring and summer and 0.9 m/s in the autumn and winter. High wind speeds of up to 25 m/s can be measured during typhoons (Qian ). Surface water flow is consistent with the wind direction. When the wind speed is low, compensated flow occurs in the bottom water, and the flow direction is opposite to the surface current. Due to the W

Back in the 1970s, Mehta and other researchers studied the relationship between sediment deposition and surface shear stress through laboratory experiments and ﬁeld trials (Mehta & Partheniades ; Partheniades ; Sheng & Lick ). Lick’s () studies in the US Great Lakes also indicated that the concentration and particle size of suspended solids are related to the shear stress of the sediment–water interface. In Lake Taihu, wind is the main driving force. Wind greatly affects the hydrodynamic conditions of the water, and hydrodynamic conditions determine the power of shear stress. Thus we can establish the relationship

Annual average hydrological and meteorological element of Lake Taihu

Lake area (km2)

Catchment area (km2)

Shoreline length (km)

Average water depth (m)

2,428

36,500

405

1.89

Maximum water depth (m)

Volume (m3)

Annual precipitation (mm)

Annual evaporation (mm)

2.8

44 × 108

1,000–1,300

1,000–1,100

Highest water level (m.a.s.l.)

Lowest water level (m.a.s.l.)

Annual sediment income (kg)

Annual sediment outcome (kg)

5.08

2.02

44 × 107

10 × 107

Average temperature ( C)

Highest temperature ( C)

Lowest temperature ( C)

Sunshine hours (h)

15.99

20.2

12.68

1,974.6

Note: The term ‘m.a.s.l.’ means ‘metres above sea level’.

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temporal and spatial variation of wind speed and direction, the flow field is unstable. But if the spatial distribution is more consistent and continues for a certain time, a clear circulation flow field can form with an average flow velocity of about 10 cm/s and a maximum of about 30 cm/s (Qin et al. ). The spatial distribution of nitrogen and phosphorus in Lake Taihu is uneven. The total nitrogen (TN) concentration varies from 0.5 to 7 mg/L, and total phosphorus (TP) concentration varies between 0.02 and 0.35 mg/L. The water quality of Meiliang Bay and Zhushan Bay is relatively poor; TN concentration is higher than 4 mg/L while TP concentration is higher than 0.15 mg/L. The water quality of East Taihu Lake is best with TN concentration less than 2.2 mg/L and TP concentration less than 0.08 mg/L. TN concentration trend for different areas of the lake is II > I > III > IV > VII > V > VI (see Figure 1 for lake areas), and TP concentration trend is II > I > IV > III > VII > V > VI (Deng et al. ). The sediment is mainly distributed in Zhushan Bay, Meiliang Bay, the west coast area and the center of the lake. Mud covers 47.5% of the lake bottom area. Sediment thickness of Meiliang Bay, Zhushan Bay, the coastal area, the north shore, the lake center, and coasts of Xishan Island are, respectively, 1.2, 0.5, 0.3, 0.4, and 0.15 m. TN content of the sediment ranges from 450 to 5,200 mg/kg but from 800 to 2,300 mg/kg in Meiliang Bay. And TP content ranges from 120 to 1,400 mg/kg but from 200 to 600 mg/kg in Meiliang Bay (Luo et al. ; Wang et al. ). In this study, the monitoring site and the in situ sediment sampling site are both site A (Figure 1; 31 250 45.14″ N, 120 70 44.85″E) in Meiliang Bay, which is severely polluted by industrial wastewater and ship transportation. Three-dimensional flow velocity about 2 cm above the water–sediment interface and wind speed were continuously measured during the field observation from July 8–10 and from October 13–15, 2013. The flow velocity was measured with an ADV and the wind speed and direction were measured with a portable aerovane.

Water Science & Technology

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Figure 1

Map of Lake Taihu with the sampling and monitoring site (site A).

Figure 2

The schematic diagram of the cylindrical simulator.

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Sediment resuspension simulate The main components of the cylindrical simulator (Figure 2) were a pump to force a vertical compensation current, a motor, a bearing turntable, two paddles to force a horizontal circular current, and a container that sediment collected into. The test section was made of clear acrylic or polycarbonate so that the sediment–water interactions could be observed. Sediment thickness was 20 cm with 80–100 cm of overlying water. Both sediment and overlying water were taken from site A and transferred to the laboratory within a few hours.

During the experiment, the average ﬂow velocity about 2 cm above the water–sediment interface was carefully controlled. The ﬂow velocity was measured with the ADV. Water was sampled through the sampling ports after the

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simulator operated at stable conditions for 30 min. About 10 test groups at a range of flow velocities were performed during which TN, TP, and SS concentrations were monitored. Measurement of SSC, TN, and TP The membrane filtration method was used to analyze the SSC in the water samples. First, the water samples were shaken thoroughly and then filtered by glass microfiber filter. Second, the filtered samples were dried at 102– 105 C for 4 h while the blank filters were dried for 2 h before filtration. Third, the samples were weighed with a precision of 0.00001 g after the sediment-laden filters were cooled to room temperature. Finally, SSC were calculated by subtracting the weight of the dry glass microfiber filter before filtration from that after the filtration. The concentration of TN and TP were measured using an automatic water quality analyzer (AA3, SEAL, Germany). Before measurement, the water samples were filtered using a water-circulation multifunction vacuum pump (SHB-III, Shiding, China). W

RESULTS AND DISCUSSION The relationship between wind speed and hydrodynamic situations The wind speed of Lake Taihu may vary from 0 to 10 m/s, with an average speed of 4.3 m/s in spring and summer and 0.9 m/s in autumn and winter (Qian ). During the field observation, the average wind speed varied from 0.2 to 3.2 m/s, while the mean flow velocity varied from 0.69 to 3.53 cm/s. Figure 3 shows the significant power relationship between average wind speed (v) and mean flow velocity (ū )

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with three different formulas. The quadratic formula (R 2 ¼ 0.969) was ū ¼ 0.262v 2 þ 0.041v þ 0.733, the liner formula (R 2 ¼ 0.914) was ū ¼ 0.907v þ 0.215, and the power formula (R 2 ¼ 0.826) was ū ¼ 1.256v 0.593. These results indicate a positive correlation between wind speed and flow velocity, and that wind forcing induces hydrodynamic conditions. The power formula (ū ¼ a•vb; a and b are constants) is often adopted (Hu et al. ; Qian et al. ). But in this study, the quadratic formula may be more suitable. The quadratic formula and liner formula differ in the zero position from the power formula. These differences may be due to the fact that the water was driven by mechanical power in simulator systems. If the mechanical power stops, the water will stop flowing and therefore there will be a zero velocity. But in Lake Taihu, many factors other than wind cause the water to flow, such as temperature and topography. When the wind speed was smaller than 0.5 m/s, the mean flow velocity was stable. In order to verify the applicability of the experimentally obtained equation, some comparisons are shown in Table 2. The quadratic formula result was closer to reference studies (Qian et al. ). In most cases, the wind speed of Lake Taihu varied between 0 and 10 m/s. Within this range, the quadratic formula is more suitable for the field and the power form is more suitable in simulator systems due to different driving conditions. The concentration and particle size of suspended solids are related to the shear stress of the sediment–water interface (Lick ). Shear stress is an important indicator for sediment resuspension flux characterization, and it is closely associated with the flow velocity. It is calculated as follows: τ ¼ ρu 2 ;

(z) u z u ¼ μ u

(1)

where τ is the shear stress (N/m2); ρ is the density of pure water (kg/m3); u* is the friction velocity (cm/s); ū (z) is the velocity at each vertical depth (cm/s); z is the vertical

Table 2

Comparisons of correspondence between wind speed and mean ﬂow velocity ū (cm/s)

v (m/s)

2.4 3.0

Figure 3

Relationship between wind speed and mean ﬂow velocity.

τ (10 3 N=m2 )

Ref. (Qian et al. 2011)

Quadratic

Linear

Power

7.9

2.2

2.3

2.4

2.1

9.5

3.4

3.2

2.9

2.4

21.5

8.2

7.5

4.8

3.3

6.5

34.2

12.9

12.1

6.1

3.8

76.4

32.2

27.3

9.3

4.9

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The regularity of wind-induced sediment resuspension

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coordinate (m); and μ is the kinematic viscosity coefﬁcient of pure water (pa s). Using Equation (1), we calculated the shear stress under different mean ﬂow velocities (Table 2) and then established the relationship between wind speed and shear stress.

and some later, and the motion could be observed in various parts of the bed in certain periods. Based on Equation (1), we were able calculate the shear stress under different ﬂow velocities (Table 3) and then SSC was calculated under different levels of shear stress.

The relationship between hydrodynamic situations and SSC

The relationship between wind speed and sediment resuspension

Due to cohesive forces and flow sweeping, sediment moved initially into groups of tens or hundreds of particles and were then transported into the water column in the form of discrete particles, leaving flaky traces behind on the bed. A remarkable increase of SSC during the experiment process was observed. There was a positive correlation between hydrodynamic forces and the resuspension rate of suspended sediment across the sediment–water interface. Figure 4 shows the relationship between SSC and flow velocity (ū ) in the cylindrical simulation experiment. The liner formula (R 2 ¼ 0.952) was SSC ¼ 10.46ū – 12.51, and the power formula (R 2 ¼ 0.982) was SSC ¼ 5.09ū 1.228. Both formulas had high coefficients of variation, but the linear formula was not suitable when ū was less than 1.2 cm/s (SSC < 0). According to the experiment, the incipient sediment motion can be generally divided into three states: first, the sediment is about to move (flow velocity ranged from 0 to 1 cm/s); second, only a small quantity of sediment is set in motion (flow velocity ranged from 1 to 4 cm/s); and third, all sediments are in motion (flow velocity ranged from 4 to 14 cm/s). In the first state, the sediments were basically motionless except for a few protrusive particles. The second referred to the state in which only a small quantity of particles could be observed in motion. The third state was one in which all the sediments were set in motion, but not all at the same time, some earlier

Through shear stress, we established links between our ﬁeld observations and the indoor experiment. We determined the relationship between wind speed and sediment resuspension. Sediment resuspension rate was calculated as follows: 2 n c0 ) þ r ¼ 4V(c

Relationship between ﬂow velocity and SSC.

Vi (c j 1 ca )5=(A t)

(2)

where r is the sediment resuspension rate (mg/(m2 d)); V is the volume of water sample in the simulator (L); cn is the nutrient concentration in water at the nth sampling (mg/L); c0 is the initial nutrient concentration (mg/L); Vi is the sample volume (L); cj 1 is the nutrient concentration in water at the ( j–1) sampling (mg/L); ca is the nutrient concentration after adding original water (mg/L); t is the release time (d); and A is the area of the water–sediment interface (m2). Sediment resuspension induced by wind is a usual phenomenon which often occurs when the bottom stress is sufﬁciently large to entrain material from the bed. Some researchers have used the power equation as a mathematical function linking SSC with wind speed in the ﬁeldworks

u (cm/s)

j¼1

Table 3

Figure 4

n X

Shear stress and SSC under different ﬂow velocity in laboratory simulation

τ (10 3 N/m2)

SSC (mg/L)

0.5

4.9

2.2

0.8

7.9

3.9

1.0

9.9

5.1

2.0

20.0

11.9

3.0

29.8

19.6

4.0

40.2

27.9

6.0

60.1

46.0

8.0

82.9

65.4

10.0

102.1

86.0

14.0

148.7

130.1

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The regularity of wind-induced sediment resuspension

(Qian et al. ; Qiao et al. ). Other expressions, such as the exponential function (You et al. ; Hu et al. ), were determined with previous laboratory simulation. The obtained expression in this work is more similar to field observations in shallow lakes. The average resuspension rate in the overlying water was highly correlated with wind speed (R 2 ¼ 0.99), which suggests a good reliability in calculating the average resuspension rate caused by wind. When the flow velocity was in the range of 0–1 cm/s (at wind speeds in the range of 0–3 m/s), sediment resuspension was almost negligible (Figure 5). When the flow velocity increased to 4 cm/s (wind speed: 7 m/s), fine particles were suspended and the water became distinctly turbid. This phenomenon indicated that the shear stress caused by the wind had reached the incipient value. According to previous observations of SSC in Lake Taihu (Qin et al. ; Hu et al. ), the critical wind speed which caused sediment resuspension was about 5–6.5 m/s. When the wind speed increased from 8 to 14 m/s, much more sediment was suspended. The phenomenon of sediment resuspension in this experiment is similar to that described by Li (). Additionally, the average resuspension rates reached 200, 1,000, and 4,400 g/(m2 d) when the wind speed was about 3, 7, and 14 m/s, respectively. This result is similar to that described by Hu et al. () who found the sediment resuspension rate was in the range of 0– 500 g/(m2 d) in spring (wind speed: 2–6 m/s), 0–1,000 g/ (m2 d) in summer (wind speed: 2–8 m/s), 0–200 g/(m2 d) in autumn (wind speed: 2–5 m/s), and 0–1,200 g/(m2 d) in winter (wind speed: 2–6 m/s). Figure 6 shows the relationship between wind speed and resuspension rates for TN and TP. The correlation of TP release rates and wind speed was strong (R 2 ¼ 0.934), but the correlation with TN was less so (R 2 ¼ 0.831). When the wind speed increased

Figure 5

Relationship between wind speed and sediment resuspension rate.

Water Science & Technology

Figure 6

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Relationship between wind speed and TN and TP resuspension rates.

to 7 m/s, the TP release rate increased to 1.1 g/(m2 d) and TN release rate reached 18.1 g/(m2 d).

CONCLUSIONS Frequent resuspension of sediments is recognized as an important process in large shallow lakes, impeding the recovery of eutrophic lakes. This paper presents field observations and a simulation experiment to study the characteristics of sediment resuspension. Our results reveal a quantitative relationship between wind speed and sediment resuspension. Our findings can be summarized as follows. (1) Flow velocity increased with wind speed. A quadratic, linear, and power relationship can be used to express the relationship between wind speed and flow velocity observed in the field. After comparison, a quadratic relationship most closely matches the results in this study. In Lake Taihu, even with no wind there is always a small flow at the bottom of the lake. This explains why this relationship in the field differs from the experimental results. (2) The SSC increased with increase of flow velocity. Compared to the linear relationship, the power formula better expressed the relationship between sediment resuspension and flow velocity wind speeds in the range of 0–14 m/s. (3) The average resuspension rate in overlying water was strongly correlated with wind speed (R 2 ¼ 0.99), which suggests wind speed can be used to accurately calculate average resuspension rates. These two factors can be expressed by r ¼ 20.72v 2.034. The critical wind speed for sediment resuspension is about 7 m/s. Under this condition, the average resuspension rate was 1,000 g/(m2 d),

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the TP release rate was 1.1 g/(m2 d), and the TN release rate was 18.1 g/(m2 d). The relationship formula, the critical wind speed and the sediment resuspension rate in this experiment are all similar to reference researches. Compared to other simulation methods of sediment resuspension, the cylindrical simulator provides more realistic results. The obtained results provided estimated sediment resuspension rates under different wind conditions. Due to the experimental scale and other interfering factors, the accuracy of the parameters derived from the simulator should be improved through further research.

ACKNOWLEDGEMENTS This work was ﬁnancially supported by the Chinese National Science Foundation (51179053) and the National Science and Technology Major Project of Water Pollution Control and Treatment (2012ZX07101-010, 2012ZX0706007, 2012ZX07506-002, 2012ZX07506-006, and 2012ZX07101001-05). We would like to thank Hohai University, Nanjing Normal University for the experimental platform, and all the people who provided help in our research process.

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Qiao, L. L., Wang, Y. Z., Li, G. X., Deng, S. G., Liu, Y. & Mu, L.  Distribution of suspended particulate matter in the northern Bohai Bay in summer and its relation with thermocline. Estuarine Coastal and Shelf Science 93 (3), 212–219. Qin, B. Q.  Lake eutrophication: control countermeasures and recycling exploitation. Ecological Engineering 35, 1569–1573. Qin, B. Q., Hu, W. P., Chen, W. M., Fan, C. X., Ji, J., Chen, Y. W., Gao, X. Y., Yang, L. Y., Gao, G., Huang, W. Y., Jiang, J. H., Zhang, S., Liu, Y. B. & Zhou, Z. Y.  Studies on the hydrodynamic processes and related factors in Meiliang Bay, Northern Taihu Lake, China. Journal of Lake Sciences 12 (4), 327–334 (in Chinese). Qin, B. Q., Hu, W. P., Gao, G., Luo, L. C. & Zhang, J. S.  Dynamic mechanism and conceptual model of inner source release of sediments suspension from Taihu Lake. Chinese Science Bulletin 48 (17), 1822–1831. Sheng, Y. P. & Lick, W.  The transport and resuspension of sediments in a shallow lake. Journal of Geophysical Research 84 (4), 1809–1826. Wan, J., Wang, Z. & Qian, S. Q.  Distribution of bioavailable phosphorus between overlying water and SPM under abrupt expansion condition. Journal of Hydrodynamics 23 (3), 398–406. Wang, P., Lu, S. Y., Wang, D. W., Xu, M. S., Gan, S. & Jin, X. C.  Nitrogen, phosphorous and organic matter spatial distribution characteristics and their pollution status evaluation of sediments nutrients in lakeside zones of Taihu Lake. China Environmental Science 32 (4), 703–709 (in Chinese).

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Xiao, L., Yang, L. Y., Zhang, Y., Gu, Y. F., Jiang, L. J. & Qin, B. Q.  Solid state fermentation of aquatic macrophytes for crude protein extraction. Ecological Engineering 35, 1668–1676. Ye, C., Yu, H. C., Kong, H. N., Song, X. F., Gou, G. Y., Xu, Q. J. & Liu, J.  Community collocation of four submerged macrophytes on two kinds of sediments in Lake Taihu, China. Ecological Engineering 35, 1656–1663. You, B. S., Zhong, J. C., Fan, C. X., Wang, T. C., Zhang, L. & Ding, S. M.  Effects of hydrodynamics processes on phosphorus fluxes from sediment in large, shallow Taihu Lake. Journal of Environmental Science 19 (9), 1055–1060. Zhai, S. J., Hu, W. P. & Zhu, Z. C.  Ecological impacts of water transfers on Lake Taihu from the Yangtze River, China. Ecological Engineering 36, 406–420. Zhao, L. L., Zhu, G. W., Chen, Y. F., Li, W., Zhu, M. Y., Yao, X. & Cai, L. L.  Thermal stratification and its influence factors in a large-sized and shallow Lake Taihu. Advances in Water Science 22 (6), 844–850 (in Chinese). Zhao, L. L., Zhu, G. W. & Xu, H.  Spatial distribution of the physicochemical parameter stratification in Meiliang Bay, Lake Taihu, China. Research of Environmental Sciences 26 (7), 721–727 (in Chinese). Zhu, G. W., Qin, B. Q. & Gao, G.  Direct evidence of phosphorus outbreak release from sediment to overlying water in a large shallow lake caused by strong wind wave disturbance. Chinese Science Bulletin 50, 577–582.

First received 28 October 2013; accepted in revised form 23 April 2014. Available online 6 May 2014

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A comparative study on the degradation of gallic acid by Aspergillus oryzae and Phanerochaete chrysosporium Danzhao Guo, Zhicai Zhang, Dan Liu, Huihua Zheng, Hui Chen and Keping Chen

ABSTRACT Recently, as an emerging persistent dissolved organic pollutant (DOP), gallic acid (GA) and its efﬁcient decomposition methods have received global attention. The present work aimed to compare the effect of Aspergillus oryzae 5992 and Phanerochaete chrysosporium 40719 on degradation of different concentrations of GA. The A. oryzae grew well and achieved a GA removal rate up to 99% in media containing 1–4% GA, much higher than P. chrysosporium. The activity of laccase and lignin peroxidase excreted by A. oryzae was higher than that by P. chrysosporium in the presence of GA. Based on the results of high-performance liquid chromatography–electrospray ionization–mass spectrometry, three relevant intermediate metabolites were determined as progallin A, methyl gallate, and pyrogallic acid, implying that A. oryzae could not degrade GA unless the carboxyl in the molecule was protected or removed. In view of the ability of A. oryzae to accommodate a high concentration of GA and achieve a high removal rate, as well as the signiﬁcantly different enzyme activities involved in GA degradation and the underlying mechanisms between the two fungal strains, A. oryzae is proven to be a superior strain for the degradation of DOP. Key words

| Aspergillus oryzae, degradation, dissolved organic pollution, gallic acid, removal rate

Danzhao Guo Zhicai Zhang Dan Liu School of Food and Biological Engineering, Jiangsu University, Zhenjiang 212013, China Zhicai Zhang (corresponding author) Institute of Agro-Production Processing Engineering, Jiangsu University, Xuefu Road 301, Zhenjiang 212013, China E-mail: zhangtamei@163.com Huihua Zheng Hui Chen Jiangsu Alphay Bio-technology Co. Ltd, Nantong 226009, China Keping Chen Institute of Life Sciences, Jiangsu University, Zhenjiang 212013, China

INTRODUCTION The main dissolved organic pollutant (DOP) in natural waters is natural organic acids (NOAs), 15–30% of which are hydrophilic small molecules. Gallic acid (GA) is one of the widely distributed hydrophilic small-molecule NOAs (Aiken et al. ; Wang et al. a, b; Chang et al. ). GA is also an important raw material used in the productions of food, household chemicals, and pharmaceuticals, the waste waters of which all contain a certain amount of GA. GA in the water causes problems as follows: (1) affects the water color, taste, and smell; (2) generates disinfection byproducts during chlorine disinfection, e.g., chloroform and haloacetic acids, which can damage organisms, especially human liver, kidney, and central nervous system (Christense et al. ), and have carcinogenic, teratogenic, and mutagenic effects too (Wang et al. ); (3) easily cause membrane fouling during the subsequent process in advanced water treatment; (4) decomposes in water into other organic acids, biogas, etc., consuming water-dissolved oxygen, resulting in hypoxia-induced death of ﬁsh, shrimp, and other aquatic doi: 10.2166/wst.2014.213

organisms; and (5) leads to increased water chemical oxygen demand, etc., resulting in a series of aquatic ecological and environmental problems (Hayatsu et al. ; Lee et al. ; Boye et al. ). Therefore, to efficiently degrade GA has become a hotspot in water pollution control. Current techniques to remove DOP from water include flocculation, adsorption, nanofiltration, and biological method (Cheng et al. ; He et al. ; Williams & Pirbazari, ; Lin et al. ). Flocculation works better for the removal of high molecular weight DOP (MW > 50,000 Da) but has a low removal rate for low molecular weight, highly water-soluble organics, and the safety of using aluminum-based and polyacrylamide flocculants has been questioned too (Wang et al. a, b). Although nanofiltration has a high removal rate for DOP in drinking water, it is less effective in depleting certain hydrophilic organic small molecules, and problems such as easy membrane fouling and congestion seriously limit its application (Figaro et al. ). Biological method can remove 30–50% of the

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DOP, without using chemical substances; it thus becomes a promising approach. Our previous study has shown that Aspergillus oryzae CGMCC5992 can effectively remove organics from vinasse in the distillery industry (Zhang et al. ). Since vinasse contains a large amount of phenols (Maestro-Durin et al. ), it is thus of interest to study whether A. oryzae CGMCC5992 can degrade phenols. Phanerochaete chrysosporium was believed to be the strongest strain in degrading aromatic compounds in wastewater (Renganathan et al. ; Huynh et al. ). Therefore, this study investigated the ability of A. oryzae to degrade phenols using GA as substrate, as well as the related pathways, with the use of P. chrysosporium 40719 as control strain.

MATERIAL AND METHODS The fungal strains A. oryzae CGMCC5992 isolated from the sludge of the Yudai River in the grounds of Jiangsu University and P. chrysosporium 40719 purchased from the China Center of Industrial Culture Collection were used in this study. Inoculation and degradation of GA A. oryzae CGMCC5992 and P. chrysosporium 40719 were both handled as below. In total, 1 × 106 spores were inoculated aseptically into a 250 mL Erlenmeyer flask with 100 mL wheat bran dextrose medium containing 20 g/L wheat bran and 20 g/L glucose and were cultured on a rotary shaker at 150 rpm, 32 C for 1–2 days. The seed culture (10 mL) was then inoculated into other 250 mL Erlenmeyer flasks containing 100 mL basal fermentation medium with the following components (g/L): glucose 13.46, NaH2PO4 24.52, CaCl2 2.18, MgSO4 9.89, and NH4Cl 4.68. Then 10, 20, 30, and 40 g/L of GA, respectively, were added to the flasks and cultured at 150 rpm, 32 C for 8 days. W

Sample preparation During the process of fermentation, the fermentation broth was sampled every 24 h and filtered with carbasius. The filter residue was analyzed for biomass, and the filtrate was centrifuged at 5,000 rpm, 4 C for 5 min to collect the supernatant, which was stored in a refrigerator at 20 C and used to analyze for GA content and enzyme activities.

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Analytical methods Analysis of biomass The fungal growth was determined by measuring the dry weight of the cells. The fermentation broth was ﬁltered with sterile carbasus; the residue was collected and washed until colorless, and as dried at 105 C to constant weight, and the weight then was designated as biomass. W

Analysis of GA content The GA in the supernatant was analyzed by a highperformance liquid chromatography (HPLC) system (Himadzu LC-20 AT Pump, N2000 station; USA), equipped with a UV detector. The analysis was performed using a RT C18 column (4.6 mm × 250 mm × 5 μm, pore size 100 Å; Australia). The mobile phase was methanol: water containing 0.1% phosphonic acid (15:85, v/v) at a constant flow rate of 0.8 mL/min. GA was detected by absorbance at 270 nm wavelength. The GA peak in samples was confirmed by spiking of pure GA, and the content of GA was calculated by external standard method. Identification of the metabolite of GA by A. oryzae using high-resolution liquid chromatography (LC)–mass spectrometry (MS) and LC-MS2 LC analysis was performed on an Agilent C-18 column (5 μm, 150 × 2.1 mm, water; Milford, MA, USA), using a high-pressure binary pump, Waters 996 diode array detector, and Agilent 1200 Series autosampler. The elution was at 0.3 mL/min with a linear gradient of acetonitrile (A) and water (B), each containing 0.1% formic acid. The gradient was from 5 to 100% A over 3 min, then to 30% A at 20 min, to 100% A at 25 min (1 min hold), followed by 5% A with a 5 min hold to equilibrate the column. The HPLC system was coupled to a Waters Platform ZMD 4000 ion trap mass spectrometer (USA) operated in the mode of full-scan negative ion electrospray ionization (ESI) (m/z 500–800). The ESI was operated in negative mode; the voltage of sample cone was at 20 V, and the collision energy was at 5 eV. Enzyme activity determination

Lignin peroxidase (LiP) activity was determined according to Tien & Kirk (). The activity of manganese peroxidase

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(MnP) was measured at 240 nm using Mn(II) as substrate (Kuwahara et al. ). Laccase activity was determined spectrophotometrically at 420 nm using guaiacol as substrate (Jiang et al. ). Statistical analysis The data of biomass and enzyme activity tests were statistically analyzed by SPSS 13.0, and all results were presented as mean of the triplicates ± standard deviation. When necessary, the results were analyzed by one-way analysis of variance.

RESULTS AND DISCUSSION Biomass determination As shown in Figure 1, the biomass of A. oryzae increased with the increase of GA concentration tested in this study during 6 days of culture, but P. chrysosporium grew poorly when the concentration of GA was over 2% (Figure 1(b)), indicating that A. oryzae 5992 was more tolerant to GA. Comparison of the GA removal rate between A. oryzae 5992 and P. chrysosporium 40719 The degradation of different initial concentrations of GA by the two strains is shown in Figure 2. The removal rates of GA by the two strains at the concentrations of 1 and 2% were obviously higher than those at the concentrations of 3 and 4% (Figures 2(a) and (b)), indicating that a lower concentration of GA was more easily degraded by the two strains. The GA removal rate of A. oryzae 5992 was obviously higher than that of P. chrysosporium 40719

Figure 1

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under each concentration of GA (Figures 2(a) and (b)), which was probably related to the higher biomass of A. oryzae 5992. The GA removal rate of P. chrysosporium reached 99% only with GA concentration lower than 2% (Figure 2(b)), whereas the removal rate of A. oryzae quickly increased in a linear manner and reached 99% with all the tested GA concentrations (Figure 2(a)). These results indicated that the ability of A. oryzae 5992 to degrade GA was stronger than that of P. chrysosporium, especially under a higher concentration of GA. Enzyme activity determination To reveal the underlying mechanism of the differential abilities in degrading GA between A. oryzae 5992 and, P. chrysosporium 40719, the activities of LiP, MnP, and laccase, which are closely related to aromatic compound degradation, were analyzed. The laccase activity of A. oryzae 5992 rapidly increased with the increase of GA concentration in the initial 4 days and reached peak value on the sixth day (Figure 3(a)). The laccase activity of P. chrysosporium 40719 was detected after 2 days of culture in the medium containing 2% GA (Figure 3(a)). Our results showed that A. oryzae 5992 secreted more laccase compared with P. chrysosporium 40719. The LiP activity was detected in the A. oryzae 5992 culture on the ﬁrst day of cultivation (Figure 3(a)), which might be derived from the seed culture. Figure 3(c) shows that although LiP activity of A. oryzae 5992 was not stable and ﬂuctuated during the whole fermentation, it was obviously higher in the medium containing 3 and 4% GA than those containing 1 and 2% GA. The LiP activity of P. chrysosporium 40719 also appeared unsteady (Figure 3(d)) and was lower than that of A. oryzae 5992 at each gradient of GA.

The biomass of A. oryzae (a) and P. chrysosporium (b). ♦: 1% gallic acid; ▪: 2% gallic acid; ▴: 3% gallic acid; ○: 4% gallic acid.

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Figure 2

Comparison of the removal rate of gallic acid by A. oryzae (a) and P. chrysosporium (b). ♦: 1% gallic acid; ▪: 2% gallic acid; ▴: 3% gallic acid; ○: 4% gallic acid.

Figure 3

The results of enzyme activities of laccase produced by A. oryzae (a); laccase produced by P. chrysosporium (b); LiP produced by A. oryzae (c); LiP produced by P. chrysosporium (d); MnP produced by A. oryzae (e); MnP produced by P. chrysosporium (f). ♦: 1% gallic acid; ▪: 2% gallic acid; ▴: 3% gallic acid; ○: 4% gallic acid.

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The MnP activity of A. oryzae peaked in the medium containing 1% GA on the second day (Figure 3(e)), which was far higher than that in the medium containing other concentrations of GA. The MnP activity from P. chrysosporium 40719 peaked on the ﬁfth day in the medium containing 1, 2, and 3% GA, respectively, and the maximum activity (615.38 U/L) was obtained in the medium supplied with 2% GA (Figure 3(f)). Many studies have proven that MnP is hardly synthesized by P. chrysosporium in a Mn2þ-free medium (Rothschild et al. ; Gold et al. ; Hofrichter ). The wheat bran used to prepare seed culture in this study contained Mn2þ, which might have induced the synthesis of MnP during the preparation of seed culture. MnP, LiP, and laccase are the three key enzymes involved in the degradation of aromatic compounds (Hatakka ). Laccase is a polyphenoloxidase that oxidizes phenol, polyphenol, pyrocatechol, and hydroquinol. Moreover, laccase is also capable of removing the methoxy group from the aromatic ring (Gianfreda et al. ; Kudanga et al. ). While LiP has the ability to non-speciﬁcally catalyze general types of substrates, including phenols, non-phenol aromatic compound, azodye, and anthraquinone dye, MnP is a kind of ectoenzyme relying on Mn(II)

Figure 4

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and H2O2 to catalyze phenol substrates (Evans et al. ; Susana et al. ; Tsukihara et al. ; Ren et al. ). A. oryzae 5992 produced a higher amount of active LiP and laccase but lower MnP compared with P. chrysosporium 40719, suggesting that the degradation of GA by A. oryzae 5992 relies mainly on LiP and laccase. Analysis of the intermediate metabolites during GA degradation by A. oryzae 5992 The metabolites in the GA-containing fermentation broth of A. oryzae at different stages were qualitatively analyzed with LC-ESI-MS. In the TIC spectra of day 0 sample, only one peak at 3.46 min was observed, and based on the quasimolecular ion m/z 169 ([M-H] ) and a fragment ion at m/z 124 in the mass spectrum of the peak, it was deduced as the peak of GA. In the TIC (total ion current) spectra of samples on the fourth day and the eighth day, four peaks (retention time: 3.43, 3.551, 5.562, and 13.224 min, respectively) were found to be related to GA metabolism, and the peak at 3.43 min was presumed as GA. The quasimolecular ion and fragment ion in the one-grade mass spectrum of these peaks are shown in Figure 4(a). In the

Diagram of the process of determination of the peak (retention time ¼ 13.224) (a) and the degradation pathway of gallic acid by A. oryzae (b).

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one-grade mass spectrum of the peak at 13.224 min, the quasimolecular ion m/z 197 ([M-H] ) and a fragment ion at m/z 169 and m/z 124 were observed, pointing to the possibility of progallin A, since lack of the fragment ion at m/z 153, m/z 182, or m/z 138 in the one-grade mass spectrum (Figure 4 (a)) ruled out the possible existence of an ethylated or methylated product of GA. Similarly, the peak at 5.562 min contained the quasi-molecular ion m/z 183, and a fragment ion at m/z 169 was recognized as methyl gallate, and the peak at 3.551 min containing quasi-molecular ion m/z 125 was pyrogallic acid. According to these data, the intermediate degradation metabolites of GA by A. oryzae included progallin A, methyl gallate, and pyrogallic acid, while other intermediated metabolites were almost undetectable. Hypothetical degradation pathway of GA by A. oryzae 5992 To further elucidate the producing order of progallin A, methyl gallate, and pyrogallic acid during the degradation process, these intermediates instead of GA were added into the medium, respectively, and the fermentation broth was analyzed by LC-ESI-MS. When progallin A or methyl gallate was added into the medium, pyrogallic acid could be detected but GA could not be detected. When pyrogallic acid was added into the medium, GA, progallin A, and methyl gallate were not detected. These results indicated that the reactions from GA to progallin A, methyl gallate, and pyrogallic acid, and from progallin A and methyl gallate to pyrogallic acid were irreversible (Figure 4(b)). In all these reactions, the fact that no intermediate metabolites other than progallin A, methyl gallate, and pyrogallic acid were detected indicated that the subsequent reaction with these intermediate metabolites was ring-opening, which was possibly a rate-limiting reaction in GA biodegradation (Osono et al. ). LiP, MnP, and laccase have been proven to be involved in the ring-opening reaction (Pérez et al. ) that produces unsaturated fatty acid, which can then be oxidized into CO2 and H2O through β-oxidation and tricarboxylic acid cycle pathways. Because these subsequent reactions are conducted in the interior of mitochondria, the intermediate metabolites are impossible to detect. Based on this, a metabolic pathway of A. oryzae 5992 degrading GA is depicted in Figure 4(b). Although we did not afﬁrm whether the pyrogallic acid was derived from GA, progallin A, or methyl gallate, and whether the ring-opening reaction occurred with pyrogallic acid, progallin A, or methyl gallate in this study, one thing certain was that A. oryzae was unable to open the ring in

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GA unless the carboxyl of the molecule was protected or removed. It is well known that if a phenolic compound contains methyl group, its degradation by white rot fungi (such as P. chrysosporium) only occurs after the removal of the methyl group; however, there is no special requirement regarding the carboxyl in the molecule. This may be the most remarkable difference in the mechanism of GA degradation between the two strains in this study.

CONCLUSION The results of biomass, removal rate of GA, enzymatic activities, and pathways involved in the degradation of different concentrations of GA in this study revealed that the signiﬁcant differences in GA degradation between the two strains are the following: (1) A. oryzae 5992 grows well in the presence of 1–4% GA and can achieve a 99% removal rate of GA; P. chrysosporium 40719 cannot grow when the concentration of GA is higher than 2%; (2) the increase in laccase activity and biomass of A. oryzae 5992 is synchronous, but the increase of laccase of P. chrysosporium 40719 lagged behind the growth of mycelia; the LiP activity in A. oryzae 5992 is far higher than that in P. chrysosporium 40719; although the maximum MnP activity shows no signiﬁcant difference between the two strains, A. oryzae 5992 reaches the maximum in a shorter time; and (3) pyrogallic acid, progallin A, and methyl gallate are the three intermediate metabolites of GA degradation, while no other metabolites are detectable, implying that there are some differences in GA degradation pathways between the two strains.

ACKNOWLEDGEMENTS This work was supported by grants from the National Natural Science Foundation of China (grant no. 31301544), Natural Science Foundation of Jiangsu Province (grant no. BK2011154), Scientiﬁc Research Initial Fund from Jiangsu University for Senior Professionals (grant no. 12JDG078) and a project funded by the Priority Academic Program Development of Jiangsu Higher Education Institutions.

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First received 26 November 2013; accepted in revised form 22 April 2014. Available online 12 May 2014