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Tweet MyronLMeters.com attempts to provide its customers with the latest in water quality research and industry updates. Find more at https://www.myronlmeters.com/. Abstract This research was conducted essentially to treat fresh peat water using a series of adsorbents. Initially, the characterization of peat water was determined and five parameters, including pH, colour, COD, turbidity, and iron ion [...]
MyronLMeters.com attempts to provide its customers with the latest in water quality research and industry updates. Find more at https://www.myronlmeters.com/.
This research was conducted essentially to treat fresh peat water using a series of adsorbents. Initially, the characterization of peat water was determined and five parameters, including pH, colour, COD, turbidity, and iron ion exhibited values that exceeded the water standard limit. There were two factors influencing the adsorption capacity such as pH, and adsorbent dosages that were observed in the batch study. The results obtained indicated that the majority of the adsorbents were very efficient in removing colour, COD, turbidity at pH range 2-4 and Fe at pH range 6-8. The optimum dosage of cationic surfactant modified zeolite (CSMZ) was found around 2 g while granular activated carbon (GAC) was exhibited at 2.5 g. In column study, serial sequence of CSMZ, GAC, and limestone showed that the optimal reduction on the 48 hours treatment were found pH = 7.78, colour = 12 TCU, turbidity = 0.23 NTU, COD = 0 mg/L, and Fe= 0.11 mg/L. Freundlich isotherm model was obtained for the best description on the adsorption mechanisms of all adsorbents.
Keywords: cationic surfactant modified zeolite, granular activated carbon, limestone, peat water
Water is essential and fundamental to all living forms and is spread over 70.9% of the earth’s surface. However, only 3% of the earth’s water is found as freshwater, of which 97% is in ice caps, glaciers and ground water (Bhatmagar & Minocha, 2006). In Malaysia, more than 90% of fresh water supply comes from rivers and streams. The demand for residential and industrial water supply has grown rapidly coupled with an increase in population and urban growth (WWF Malaysia, 2004). Water demand in affected populations such as rural areas also demands that attention is paid to providing more sustainable solutions rather than transporting bottled water (Loo et al., 2012). For this reason, it is essential to ensure availability of local sources of water supply and even develop new potential sources of water such as from peat swamp forest to overcome future water shortages.
River water surrounded by peat swamp forest is defined as peat water and is commonly available as freshwater since it has a low concentration of salinity. The previous study shows that peat swamp forest has high levels of acidity and organic material depending on its region and vegetation types (Huling et al., 2001). Under natural conditions, tropical peat lands serve as reservoirs of fresh water, moderate water levels, reduce storm-flow and maintain river flows, even in the dry season, and they buffer against saltwater intrusion (Wosten et al., 2008).
Due to the acidity and high concentration of organic material, selective treatment of peat water must be conducted prior to its use as water supply. Recently, many methods have been designed and have proven their effectiveness in treating raw water such as coagulation and flocculation (Franceschi et al., 2002; Liu et al., 2011; Syafalni et al., 2012a), absorption (Ćurković et al., 1997), filtration (Paune et al., 1998) and combining (Hidaka et al., 2003). Careful consideration of the most suitable method is important to ensure that the adsorption process is the most beneficial, economically feasible method as well as easy to operate for producing high quality of water in a particular location.
Many researchers have shown that activated carbon is an effective adsorbent for treating water with high concentrations of organic compounds (Eltekova et al., 2000; Syafalni et al., 2012b). Its usefulness derives mainly from its large micropore and mesopore volumes and the resulting high surface area (Fu & Wang, 2011). However, its high initial cost makes it less economically viable as an adsorbent. Low cost adsorbent such as zeolite nowadays has been explored for its ability in many fields especially in water treatment. Natural zeolite has negative surface charge which gives advantages in absorbing unwanted positive ions in water such heavy metal. These ions and water molecules can move within the large cavities allowing ionic exchange and reversible rehydration (Jamil et al., 2010). The effectiveness of zeolite has been improvised by modified zeolite with surfactant in order to achieve higher performance in removing organic matter (Li & Bowman, 2001). Among tested cationic surfactants, hexa-decyl-tri-methyl ammonium (HDTMA) ions adsorbed onto adsorbent surfaces are particularly useful for altering the surface charge from negative to positive (Chao & Chen, 2012). Surfactant modified zeolite has been shown to be an effective adsorbent for multiple types of contaminants (Zhaohu et al., 1999).
Zeolite is modified to improve its capability of exchanging the anion by cationic surfactants, called CSMZ. CSMZ adsorbs all major classes of water contaminants (anions, cations, organics and pathogens), thus making it reliable for a variety of water treatment applications (Bowman, 2003). Nowadays, interest in the adsorption of anions and neutral molecules by surfactant modified zeolite has increased (Zhang et al., 2002). Modification of zeolite by surfactant is commonly done by cationic or amphoteric surfactants. By introducing surfactant to the zeolite, an organic layer is developed on the external surfaces and the charge is reversed to positive (Li et al., 1998). However, the present study used zeolite that had been modified using Uniquat (QAC-50) as cationic surfactant (CSMZ) and their performance towards the removal of color, COD, turbidity and iron ion from peat water were investigated.
Four adsorbents were used in these experiments which are natural zeolite, zeolite modified by cationic surfactant, activated carbon and limestone. All adsorbents were prepared with equivalent sizes of 1.18 mm – 2.00 mm. Hydrochloric acid (HCl) and sodium hydroxide (NaOH) were used for polishing zeolite during the preparation phase and for pH adjustment of the sample. Furthermore, potassium dichromate (K2CrO7), silver sulphate (Ag2SO4), sulphuric acid (H2SO4) and mercury (II) sulphate (HgSO4) were used as digestion solution reagents and acid reagents for COD analysis. Lastly, Uniquat (QAC-50) was used as cationic surfactant to modify the zeolite.
2.1 Preparation of Surfactant Modified Zeolite
In these studies, 100 g of prewashed natural zeolite was contacted with 5.6 ml/l Uniquat (QAC-50) as cationic surfactant (CSMZ). The mixture was then stirred at room temperature for 4 hours at 300 rpm (Karadag et al., 2007). The zeolite then was filtered and washed with distilled water several times. After that, the absorbent was dried in an oven at a temperature of 105 °C for 15 hours.
2.2 est Procedures
2.2.1 Batch Studies
Serial batch studies were conducted at room temperature (28 ± 1 °C) to investigate the influence of pH and dosage for removing colour, COD, turbidity and iron ion from peat water. Shaking speed of 200 rpm for 20 minutes were fixed and operated respectively. A working volume of 150ml peat water sample was set up in 250 ml conical flasks. Preceding the batch studies, initial concentration for those parameters was determined. The optimum pH and dosage of absorbent were determined. Subsequently, the percentage of removal was finally determined, plotted, and compared.
2.2.2 Batch Column Studies
Column studies were carried out using a plastic column with dimensions: 5.4 cm diameter and 48 cm length. Three adsorbents were filled inside the column at a specific depth with the supporting layers of marbles, cotton wool, and perforated net. Total volume of 2000 ml peat water was pumped in the up flow mode from the vessel into the column by using a Masterflex peristaltic pump at a minimum flow rate of (30, 60, 90) ml/min. In this study, however, column studies were performed un-continuously (batch) due to limitations of time. All parameters related to the column design are summarized in the following Table 1.
Table 1. Column studies parameters
|Horizontal Surface Area, A
|Column volume, V
||30, 60, 90
|Surface Loading Rate, SLR= Q/A
||1.31, 2.62, 3.93
The serial sequence arrangements of adsorbents were conducted as shown in Figure 1 below. Effluent samples were collected at various time intervals, whilst maintaining room temperature, and analysed.
Figure 1. Schematic diagrams of lab-scale column studies
3. Results and Discussion
3.1 eat Water Characterization
Surface water originating from the peat swamp forest was taken from Beriah peat swamp river along the Kerian River on several occasions as the main sample. The characterization of peat water was carried out at the sampling point (in-situ measurement) using a multi-parameter probe as well as in the environmental laboratory of civil engineering, USM. Fundamentally, the characterization procedures were based on the Standard Methods for the Examination of Water and Wastewater (APHA, 1992). Table 2 represents the peat water characteristics in average value and the comparison to the standard drinking water quality in Malaysia.
Table 2. The characteristics of peat water sample from Beriah Peat Swamp Forest
|4.67 – 4.98
Thirteen parameters were successfully determined where the first six parameters, including pH, temperature, TDS, DO, conductivity, and salinity were measured at the sampling point, whilst the rest of the parameters, including colour, turbidity, COD, iron ion, Ammoniacal Nitrogen, NH3-N, Ammonia (NH3), and Ammonium (NH4+) were examined from the sample brought to the environmental laboratory on the same day.
Acidic pH of the peat water was predicted due to the composition of the surrounding peat soil itself which had been formed by decaying material possessing humic substances (Rieley, 1992). Besides that, humic substances also lead to the high organic content as humic substances are comprised of numerous oxygen containing functional group and fractions (humic acid, fulvic acids and humin) with different molecular weights which mean yielding high concentration of turbidity and COD as well as coloured water (Torresday et al., 1996). Moreover, composition of peat soil may also have an impact on the iron ion concentration of peat water (Botero et al., 2010).
From the thirteen parameters, five parameters were indicated exceeding the standard limit. These parameters were pH, colour, turbidity, COD, and iron ion that showed values of 4.67 – 4.98, 224.7 TCU, 20.8 NTU, 33.3 mg/l, and
1.24 mg/l respectively while the standard limit of these parameters are 6.5 – 9.0, 15 TCU, 5 NTU, 10 mg/l, and 0.3 mg/l accordingly.
3.2 Effect of Initial pH on the Efficiency of Colour, COD, Turbidity, and Iron Ion (Fe) Removal
Influence of initial pH on the adsorption capacity for removing colour, COD, turbidity, and iron ion were investigated.
Figure 2(a) to Figure 2(d) below, displayed the percentage removal of colour, COD, turbidity, and iron ion against pH of adsorbents respectively.
Figure 2(a) shows the maximum removal percentage of colour that was removed by natural zeolite, CSMZ, and granular activated carbon (GAC) which were 79%, 90%, 82% respectively. This adsorption is depended on the characteristic of adsorbents itself. For zeolite and CSMZ were related to the amount of cationic ions (Al3+) increased, resulting in high reaction activity and GAC was related to the adsorption capacity. It was observed that the adsorption capacity was highly dependent on the pH of the solution, and indicated that the colour removal efficiencies decreased with the increase of solution pH.
The pH of the system exerts profound influence on the adsorptive uptake of adsorbate molecules presumably due to its influence on the surface properties of the adsorbent and ionization or dissociation of the adsorbate molecule. Figure 2(b) represents the percentage removal of natural zeolite and CSMZ where they reach optimum efficiency in removing organic compound (COD) at pH 2 with efficiency of 53% and 60% respectively. Meanwhile, the highest percentage removal of COD for GAC was achieved at pH 4 with efficiency obtained about 61%. Identical trends in colour removal were exhibited in percentage removal of COD for natural zeolite, CSMZ and GAC. In fact, this result also reveals that GAC has the highest percentage removal among natural zeolite and CSMZ yet optimum in difference pH solution. Neutralization mechanism occurs in low pH makes color removal, COD removal and Turbidity removals at pH 2 are higher for most of absorbents in this process.
In Figure 2(c), percentage turbidity removal against pH for each adsorbent revealed that optimal reduction of turbidity was obtained in an acidic environment with efficiency removal of 96%, 98%, 95% for natural zeolite, CSMZ, and GAC respectively. When the pH of the solution was adjusted above pH 6 to pH 12, the tendencies of all adsorption performances were gradually decreased. Moreover, it also showed that the lowest efficiency for the three adsorbents were identified at pH 12 with percentage values removal 55%, 61%, and 59% for natural zeolite, CSMZ, and GAC respectively.
Figure 2(d) demonstrates the removal efficiencies of iron ion as a function of the influent pH. The maximum removal of iron ion was observed at pH 8 for both natural zeolite and CSMZ whereas GAC had its optimum removal at pH 6. Natural zeolite and CSMZ only yielded 73% and 62% removal efficiency while GAC had more significant removal with removal efficiency of 80% to the iron ion concentration. Further, it is evident from the graph that gradual increment of removal efficiency for natural zeolite, CSMZ, and GAC occurred when the initial pH of the solution was increased to higher values. Somehow, at pH values greater than 6 the removal efficiency of GAC reduced slightly while for natural zeolite and CSMZ the reduction occurred from pH values above 8.
3.3 Effect of Adsorbent Dosage on the Efficiency of Colour, COD, Turbidity, and Iron Ion (Fe) Removal
The effect of adsorbent dosage was studied for all adsorbents employed on colour, COD, turbidity, and iron ion removal by varying the dosage of adsorbent and keeping all other experimental conditions constant. The pH was set to acidic conditions which were most favourable in obtaining the highest removal efficiency. In this study, to find optimal adsorbent dosage of natural zeolite and CSMZ, the appropriate experiments were carried out at adsorbent dosages in the range of 0.5 g to 5.0 g while for GAC, the adsorbent dosage was varied from 0.01 g to 4.0
- The experimental results for all the adsorbents are represented by Figure 3(a) to Figure 4(d).
Figure 3. Percentage of color (a), COD (b), turbidity (c), and Fe (d) removal against pH for NZ, and CSMZ
Figure 3(a) displays the relationship between the amount of adsorbent mass (dosage) and adsorption efficiency for natural zeolite and CSMZ in terms of removing colour. The colour removal of peat water increased from about 25% to 52% with increasing adsorbent dosage of natural zeolite from 0.5 g to 3.5 g whereas for CSMZ, removal percentage increased from 41% to 53% with increasing adsorbent dosage from 0.5 g to 2.0 g. However, further increase in adsorbent dosage to 5.0 g only led to slight degradation of removal efficiency to 50% and 41% for natural zeolite and CSMZ respectively. This degradation with further increases in adsorbent dosage was due to the unsaturated adsorption active sites during the adsorption process since the adsorbates in the vessel were only shaken for 20 minutes (insufficient time). Besides, modification of zeolite by cationic surfactant had proven to have better colour removal as presented in the graph.
Percentage removal of COD against the adsorbent dosage is shown in Figure 3(b). It was observed that the highest percentage removal for both natural zeolite and CSMZ to remove COD were 51% and 59%, achieved at adsorbent dosage 3.5 g and 2.0 g respectively.
The variations in removal of turbidity of peat water at various system pH are shown in Figure 3(c). The removal rate of turbidity was highest at the adsorbent dosage of 0.5 g with 70% and 93% removal efficiency for respective natural zeolite and CSMZ. The removal rate showed a smooth downward trend with the increase in adsorbent dosage. Concurrently, the adsorption capacity gradually decreased with the increasing adsorbent dosage. The least efficient removal of turbidity was noted at dosage 5.0 g with percentage removal recorded for natural zeolite and CSMZ only 57% and 70% respectively.
Figure 3(d) demonstrates the percentage iron ion removal of natural zeolite and CSMZ with respect to their dosage. The result shows that there was a significant difference trend in iron ion adsorption efficiencies between natural zeolite and CSMZ. For natural zeolite, it was shown that the removal percentage of iron ion had increased until it reached 1.0g of dosage with 72% of removal efficiency. On the other hands, CSMZ was only able to remove about 63% of iron ion when its dosage was increased to 2.5 g. The lowest percentage removals were 47% and 57% recognized at the adsorbent dosage 5.0 g for respective natural zeolite and CSMZ.
Figure 4. Percentage of color (a), COD (b), turbidity (c), and Fe (d) removal against dosage for GAC
The result illustrated in Figure 4(a) shows the maximum removal percentage of colour for GAC at 2.5 g dosage was 62%. Moderate increment in colour removal was identified along with the addition dosage of 2.5 g whilst abatement of removal efficiency began subsequently at adsorbent dosage of 3.0 g to 4.0 g.
The results from Figure 4(b) indicated that increasing the GAC dosage would increase the efficiency in removing COD respectively. The optimum dosage was recorded at 3.0 g with 72% of removal efficiency. Meanwhile, increasing the dosage above 3.0 g exhibited a slight decrease in removal efficiency with 67% to 61% for COD removal. A better result in removing COD was also shown by GAC compared to the natural zeolite and CSMZ.
The percentage of turbidity removed by GAC in different dosages is described in Figure 4(c). The highest removal was indicated at adsorbent dosage 2.5 g with removal efficiency of 70% while the minimum removal was 52% recorded at the adsorbent dosage 0.01 g. However, starting from adsorbent dosage of 3.0 to 4.0 g, removal efficiency began to decrease to 68%, 67%, and 69% respectively.
The result of percentage removal of iron ion by GAC in peat water is presented in Figure 4(d). It was found that the rate of removal was rapid in the initial dosage between 0.01 g to 3.0 g at which the removal efficiency increased from 28% to 71% accordingly. Subsequently, a few significant changes in the rate of removal were observed. Possibly, at the beginning, the solute molecules were absorbed by the exterior surface of adsorbent particles, so the adsorption rate was rapid. However, after the optimum dose was reached, the adsorption of the exterior surface becomes saturated and thereby the molecules will need to diffuse through the pores of the adsorbent into the interior surface of the particle (Ahmad & Hameed, 2009).
3.4 Batch Column Experiment
On the first running, the column was packed with natural zeolite (1st layer), limestone (2nd layer), and GAC (3rd layer) as shown in Figure 5(a). Removal efficiency for colour, COD, turbidity, and iron ion was recognized to be increased when the contact time was increased. At the time interval 1 hour to 6 hours, however, the increment was not so significant. The removal efficiency at 1 hour treatment was 39%, 21%, 54%, 36% while at 6 hours treatment was 77%, 65%, 73%, 60% recorded for respective colour, COD, turbidity, and iron ion. Poor removal efficiency at 1 hour treatment indicated that the required time to remove all parameters were insufficient. It is evident that if the adsorption process is allowed to run for 24 hours on the column, the removal efficiency shows notable removal. Percentage removals of colour, COD, turbidity, and iron ion at 24 hours were 83%, 72%, 76%, 65% respectively. Furthermore, the highest removal for respective colour, COD, turbidity, and iron ion were obtained at 48 hours treatment with 87%, 81%, 86%, and 79% of removal efficiency.
Figure 5. Percentage removal of color, COD, turbidity, and Fe for 1st run(a), 2nd run(b), and 3rd run (c) at flowrate 30 ml/min
On the second running, the column was packed with CSMZ (1st layer), limestone (2nd layer), and GAC (3rd layer) as presented in Figure 5(b). The removal percentages of colour, COD, turbidity, and iron ion were noticed after 1 hour to be 52%, 49%, 71%, and 30% respectively. The time of contact between adsorbate and adsorbent is proven to play an important role during the uptake of pollutants from peat water samples by adsorption process. In addition, the development of charge on the adsorbent surface was governed by contact time and hence the efficiency and feasibility of an adsorbent for its use in water pollution control can also be predicted by the time taken to attain its equilibrium (Sharma, 2003). Removal efficiency of 90% for colour, 81% for COD, 91% for turbidity, and 57% for iron ion were obtained at 24 hours of contact time.
On the third running, the column was packed with a difference sequence of CSMZ (1st layer), GAC (2nd layer), and limestone (3rd layer) demonstrated in Figure 5(c). It can be seen that the adsorption of these four parameters were slightly rapid at time interval 1 hour to 6 hours treatment. Further gradual increment with the prolongation of contact time form 24 hours to 48 hours has also occurred. Observation at 1 hour treatment recorded the removal efficiency of 62%, 58%, 87%, and 48% for respective colour, COD, turbidity, and iron ion. Whereby, 6 hours treatment had yielded higher removal percentage removal of 75%, 77%, 93%, and 58% respectively for colour, COD, turbidity, and iron ion. Further removal of colour, COD, turbidity, and iron ion was recorded when the treatment was run for 24 hours which exhibited 92%, 91%, 98%, 77% of removal efficiency respectively. Prolonged time to 48 hours indeed showed better removal of colour, COD, turbidity, iron ion with percentage removal of 95%, 100%, 99%, and 89% respectively. It can be seen that the arrangement of CSMZ, GAC, and limestone has the highest removal efficiency for all parameters at the flow rate influent of 30 ml/min.
Figure 6. Percentage removal of color, COD, turbidity, and Fe against contact time for 2nd run(a) at flow rate 60 mL/min and at flowrate 90 mL/min (b)
The experimental adsorption behaviour was further seen for its adsorption capacity during 60 ml/min and 90 ml/min flow rate. In addition, the flow rate adjustment had also resulted in differences in surface loading rate in which the sample going through the surface area of adsorbent bed (horizontal surface area, A= 22.9 cm2) for 30 ml/min equals to 1.31 cm/min while the flow rate of 60ml/min equals to 2.62 cm/min, and the flow rate of 90 ml/min equals to 3.93 cm/min. The percentage removal for both flow rate adjustments of CSMZ, GAC, and limestone arrangement were exhibited in Figure 6 (a) and Figure 6 (b). Based on these Figures, lower removal efficiencies were indicated at 1 hour time interval of 6 hours of contact time. The percentage removals for both 60 ml/min and 90 ml/min flow rate at 1 hour were 57%, 56%, 80%, 38% and 49%, 58%, 61%, 35% for colour, COD, turbidity, and iron ion respectively. Subsequently, when the contact time was at 6 hours, the removal percentage were 70%, 79%, 88%, 56%, and 60%, 77%, 70%, 47%. However, the maximum removal efficiency at 48 hours for both flow rates was not much different from the 30ml/min flow rate.
3.5 Adsorption Isotherm
In the present investigation, the experimental data were tested with respect to both Freundlich and Langmuir isotherms. Based on the linearized Freundlich isotherm models for natural zeolite, CSMZ, GAC in terms of adsorptive capacity to remove colour, COD, turbidity, and iron ion, the majority of them exhibited fits for all adsorbate with regression value (R2) above 0.6, except for iron ion and turbidity for respective CSMZ, and GAC. On the other hand, the linearized Langmuir isotherm models for natural zeolite, CSMZ, GAC in terms of adsorptive capacity to remove colour, COD, turbidity, and iron ion, had exhibited fits for all adsorbate with regression value (R2) was at range of 0.242 to 0.912. The Langmuir isotherm model for all adsorption mechanisms were identified to have smaller R2 values compared to the Freundlich isotherm model. Thereby, it can be concluded that the Freundlich isotherm model was more applicable in determining the adsorption mechanisms for this study.
3.6 Peat Water Quality Post Column Treatment
Peat water treatment in column with serial sequence of natural zeolite, CSMZ, and limestone had exhibited the highest removal with percentage removal at 48 hours at 95%, 100%, 99%, and 89% for colour, COD, turbidity, and iron ion respectively. Final readings at 48 hours treatment on pH, TDS, DO, conductivity, salinity, colour, turbidity, COD, and iron ion were 7.78, 74 mg/l, 4.03 mg/l, 137 uS/cm, 0.05 ppt, 12 TCU, 0.23 NTU, 0 mg/l, and 0.11 mg/l respectively (see Table 3). These findings, on the other hand, have indicated that peat water treatment had successfully produced water which satisfied the standard drinking water quality.
Table 3. The characteristics of results of peat water treatment from Beriah Peat Swamp Forest
Note: 1. *)Malaysian standard for drinking water quality;2. NA = Not analyzed.
From the results presented in this paper, the following conclusions can be drawn:
1) The optimum removal of colour, COD, and turbidity for all adsorbents were observed to occur during acidic conditions at pH range 2 – 4 whereas for iron ion, the maximum removal was noted at pH range 6 – 8.
2) At pH 2, CSMZ yielded the highest removal for colour and turbidity with removal efficiency of 90% and 98% respectively. Meanwhile, GAC has the highest percentage removal of COD at pH 4 with removal efficiency obtained about 61% while at pH 6, GAC exhibited the best removal of iron ion with percentage removal around 80%.
3) CSMZ revealed stronger adsorptive capacity for colour, COD, and turbidity compared to natural zeolite.
4) The optimal removal was achieved for the serial sequence of CSMZ (1st layer), GAC (2nd layer), and Limestone (3rd layer) with the adsorbent media at 30 ml/min of flow rate.
5) Freundlich isotherm was more reliable to describe the adsorption mechanisms of colour, COD, turbidity, and iron ion for natural zeolite, CSMZ, and GAC.
The authors wish to acknowledge the financial support from the School of Civil Engineering, Engineering Campus, Universiti Sains Malaysia and Universiti Sains Malaysia (Short Term Grant No. 304/PAWAM/60312015).
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Wosten, J. H. M., Clymans, E., Page, S. E., Rieley, J. O., & Limin, S. H. (2008). Peat- Water interrelationships in a Tropical Peatland Ecosystem in Southeast Asia. Catena, 73, 212-224. http://dx.doi.org/10.1016/j.catena.2007.07.010
Zhang, P., Tao, X., Li, Z., & Bowman, R. S. (2002). Enhanced Perchloroethylene Reduction in Column Systems Using Surfactant Modified Zeolite/zero-valent Iron Pellets. Environmental Science and Technology, 36, 3597-3603. http://dx.doi.org/10.1021/es015816u
Modern Applied Science; Vol. 7, No. 2; 2013
ISSN 1913-1844 E-ISSN 1913-1852
Published by Canadian Center of Science and Education
S. Syafalni1, Ismail Abustan1, Aderiza Brahmana1, Siti Nor Farhana Zakaria1 & Rohana Abdullah1
1 School of Civil Engineering, Engineering Campus, Universiti Sains Malaysia, Nibong Tebal, Penang, Malaysia. Correspondence: S. Syafalni, School of Civil Engineering, Engineering Campus, Universiti Sains Malaysia,
Nibong Tebal 14300, Penang, Malaysia. E-mail: email@example.com
Received: December 3, 2012 Accepted: January 14, 2013 Online Published: January 22, 2013 doi:10.5539/mas.v7n2p39 URL: http://dx.doi.org/10.5539/mas.v7n2p39
Shared via Creative Commons Attribution 3.0 Unported license
admin 11 Apr, 2013
Tweet Myron L Meters Myron L Meters sells the most accurate, reliable conductivity instruments in the water treatment industry. You can find some of our most popular meters here: http://www.myronlmeters.com/SearchResults.asp?Search=conductivity&x=-1345&y=-145 Introduction Along with the development of more and more complex integrated models for urban water systems the need of sufficient data bases grows as well. [...]
Myron L Meters
Myron L Meters sells the most accurate, reliable conductivity instruments in the water treatment industry. You can find some of our most popular meters here:
Along with the development of more and more complex integrated models for urban water systems the need of sufficient data bases grows as well. It is even complicated to measure relevant parameters,e.g. dissolved nitrogen or COD, for their use in Waste Water Treatment Plants and sewer models to describe the influence of catchments to the receiving water.
This poster presents a method regarding the possibility of substituting an online ammonia measurement by conductivity measurements in the inflow of a Waste Water Treatment Plant . The aim was the description of the dynamics in wet weather flow through storm water events for modelling purposes.
The conductivity of an aqueous solution is the measure of its ability to conduct electricity. Responsible for that phenomenon are ions of dissolved salts. In natural and drinking water these are mainly carbonates, chlorides and sulphates of calcium, magnesium, sodium and potassium. Conducted experiences and measurements in combined sewers showed a relation between conductivity in Waste Water Treatment Plant inflow and the concentration of dissolved components, e.g. ammonia, in case of rainfall events. The data for different 3 Waste Water Treatment Plant are shown in Figure 1. Rainwater has nearly no ions that cause conductivity to be measured. Therefore, diluted wastewater flowing into the Waste Water Treatment Plant can be detected by a conductivity probe. The measure and quality of linear regression between ammonia concentration and conductivity can be found in Table 1 for all data from Figure 1.
Material and Methods
With this knowledge a simple regression-based inflow model for use in activated sludge modelling of Waste Water Treatment Plant was defined to use conductivity beside available composite samples as a measure for dynamics in ammonia concentration as one of the most dynamic measure.
Results and Discussion
For one of the considered Waste Water Treatment Plants (WWTP) the resulting quality for the inflow model is shown in Figure 2 for a time series of a week.
Furthermore, the inflow model was used as a source for a retention tank model at the inlet of another Waste Water Treatment Plant to describe the impact of different management strategies (storage or flow through) on receiving water and Waste Water Treatment Plant (Figure 3).
A long-term modelling of 9 storm water events was used to show the predictive capacity of the model. The regression parameters were fitted by an optimisation routine to get best fit for all concentrations (also for COD, not presented here). Figure 3 shows the fit for all events. A good prediction of dynamics and absolute values for ammonia can be seen.
The results of different Goodness-of-fit measures are summarized in Table 2 for both presented WWTP inflows. Especially the values for the modified Coefficient of Efficiency, as a well-known and used measure for model quality in hydrological sciences, show the degree of predicting of the used method and the usability of conductivity for description of influent dynamics to Waste Water Treatment Plant in storm water cases.
This simple and easy-to-use method is well suited for implementation in Waste Water Treatment Plant models to describe the inflow dynamics regarding a more realistic behavior e.g. for optimization of process control.
by Markus Ahnert*, Norbert Günther*, Volker Kuehn*, University of Dresden
Ahnert, M., Blumensaat, F., Langergraber, G., Alex, J., Woerner, D., Frehmann, T., Halft, N., Hobus, I., Plattes, M., Spering, V. und Winkler, S. (2007), Goodness-of-fit measures for numerical modelling in urban water management – a summary to support practical applications., paper presented at 10th IWA Specialised Conference on “Design, Operation and Economics of Large Wastewater Treatment Plants”, 9-13 September 2007, Vienna, Austria, 9-13 September 2007.
Nash, J. E. und Sutcliffe, J. V. (1970), River flow forecasting through conceptual models part I – A discussion of principles, Journal of Hydrology, 10, 282.
admin 3 Apr, 2013
TweetThe thermal conductivity enhancement of nanofluids Abstract Increasing interests have been paid to nanofluids because of the intriguing heat transfer enhancement performances presented by this kind of promising heat transfer media. We produced a series of nanofluids and measured their thermal conductivities. In this article, we discussed the measurements and the enhancements of the thermal [...]
The thermal conductivity enhancement of nanofluids
Increasing interests have been paid to nanofluids because of the intriguing heat transfer enhancement performances presented by this kind of promising heat transfer media. We produced a series of nanofluids and measured their thermal conductivities. In this article, we discussed the measurements and the enhancements of the thermal conductivity of a variety of nanofluids. The base fluids used included those that are most employed heat transfer fluids, such as deionized water (DW), ethylene glycol (EG), glycerol, silicone oil, and the binary mixture of DW and EG. Various nanoparticles (NPs) involving Al2O3 NPs with different sizes, SiC NPs with different shapes, MgO NPs, ZnO NPs, SiO2 NPs, Fe3O4 NPs, TiO2 NPs, diamond NPs, and carbon nanotubes with different pretreatments were used as additives. Our findings demonstrated that the thermal conductivity enhancements of nanofluids could be influenced by multi-faceted factors including the volume fraction of the dispersed NPs, the tested temperature, the thermal conductivity of the base fluid, the size of the dispersed NPs, the pretreatment process, and the additives of the fluids. The thermal transport mechanisms in nanofluids were further discussed, and the promising approaches for optimizing the thermal conductivity of nanofluids have been proposed.
More efficient heat transfer systems are increasingly preferred because of the accelerating miniaturization, on the one hand, and the ever-increasing heat flux, on the other. In many industrial processes, including power generation, chemical processes, heating or cooling processes, and microelectronics, heat transfer fluids such as water, mineral oil, and ethylene glycol always play vital roles. The poor heat transfer properties of these common fluids compared to most solids is a primary obstacle to the high compactness and effectiveness of heat exchangers. An innovative way of improving the thermal conductivities of working media is to suspend ultrafine metallic or nonmetallic solid powders in traditional fluids since the thermal conductivities of most solid materials are higher than those of liquids. A novel kind of heat transfer enhancement fluid, the so-called nanofluid, has been proposed to meet the demands .
“Nanofluid” is an eye-catching word in the heat transfer community nowadays. The thermal properties, including thermal conductivity, viscosity, specific heat, convective heat transfer coefficient, and critical heat flux have been studied extensively. Several elaborate and comprehensive review articles and books have addressed thermal transport properties of nanofluids [1,3-6]. Among all these properties, thermal conductivity is the first referred one, and it is believed to be the most important parameter responsible for the enhanced heat transfer. Investigations on the thermal conductivity of nanofluids have been drawing the greatest attention of the researchers. A variety of physical and chemical factors, including the volume fraction, the size, the shape, and the species of the nanoparticles (NPs), pH value and temperature of the fluids, the Brownian motion of the NPs, and the aggregation of the NPs, have been proposed to play their respective roles on the heat transfer characteristics of nanofluids [7-19]. Extensive efforts have been made to improve the thermal conductivity of nanofluids [7-19] and to elucidate the thermal transport mechanisms in nanofluids [20-23].
The authors have carried out a series of studies on the heat transfer enhancement performance of nanofluids. A variety of nanofluids have been produced by the one- or two-step method. The base fluids used include deionized water (DW), ethylene glycol (EG), glycerol, silicone oil, and the binary mixture of DW and EG (DW-EG). Al2O3 NPs with different sizes, SiC NPs with different shapes, MgO NPs, ZnO NPs, SiO2 NPs, Fe3O4 NPs, TiO2 NPs, diamond NPs (DNPs), and carbon nanotubes (CNTs) with different pretreatments have been used as additives. The thermal conductivities of these nanofluids have been measured by transient hot wire (THW) method or short hot wire (SHW) technique. In this article, the experimental results that elucidate the influencing factors for thermal conductivity enhancement of nanofluids are presented. The thermal transport mechanisms in nanofluids and promising approaches for optimizing the thermal conductivity of nanofluids are further presented.
Preparation of nanofluids
Two techniques have been applied to prepare nanofluids in our studies: two- and one-step techniques. Most of the studied nanofluids were prepared by the two-step technique. During the procedure of two-step technique, the dispersed NPs were prepared by chemical or physical methods first, and then the NPs were added into a specified base fluid, with or without pretreatment and surfactant based on the need. In the preparation of nanofluids containing metallic NPs, one-step technique was employed.
The process was quite simple in the preparation of nanofluids containing oxide NPs like Al2O3, ZnO, MgO, TiO2, and SiO2 NPs. The NPs were obtained commercially and were dispersed into a base fluid in a mixing container. The NPs were deagglomerated by intensive ultrasonication after being mixed with the base fluid, and then the suspensions were homogenized by magnetic force agitation.
Two-step method was used to prepare graphene nanofluids. The first step was to prepare graphene nanosheets. Functionalized graphene was gained through a modified Hummers method as described elsewhere . Graphene nanosheets were obtained by exfoliation of graphite in anhydrous ethanol. The product was a loose brown powder, and it had good hydrophilic nature. The graphene nanosheets could be dispersed well in polar solvents, like DW and EG, without the use of surfactant. For liquid paraffin (LP)-based nanofluid, oleylamine was used as the surfactant. The fixed quality of graphene nanosheets with different volume fractions was dispersed in the base fluids.
Severe aggregation always takes place in the as-prepared CNTs (pristine CNTs: PCNTs) because of the non-reactive surfaces, intrinsic Von der Waals forces, and very large specific surface areas, and aspect ratios . In CNT nanofluid preparations, surfactant addition is an effective way to enhance the dispersibility of CNTs [26-28]. However, surfactant molecules attaching on the surfaces of CNTs may enlarge the thermal resistance between the CNTs and the base fluid , which limits the enhancement of the effective thermal conductivity. The steps involved in the preparation of surfactant-free CNT nanofluids include (1) disentangling the nanotube entanglement and introducing hydrophilic functional groups on the surfaces of the nanotubes by chemical treatments; (2) cutting the treated CNTs (TCNTs) to optimal length by ball milling; and (3) dispersing the treated and cut CNTs into base fluids. CNTs including single-walled CNTs (SWNTs), double-walled CNTs (DWNTs), and multi-walled CNTs (MWNTs) were obtained commercially. Two chemical routes for treating CNTs were used for this study. One is oxidation with concentrated acid, and the other is mechanochemical reaction with potassium hydroxide (KOH). The detailed treatment processes have been described elsewhere [8,30].
Phase transfer method was used to prepare stable kerosene-based Fe3O4 magnetic nanofluid. The first step is to synthesize Fe3O4 NPs in water by coprecipitation. Oleic acid was added to modify the NPs. When kerosene is added to the mixture with slow stirring, the phase transfer process took place spontaneously. There was a distinct phase interface between the aqueous and kerosene. After the removal of the aqueous phase using a pipette, the kerosene-based Fe3O4 nanofluid was obtained .
Nanofluids containing copper NPs were prepared using direct chemical reduction method. Stable nanofluids were obtained with the addition of poly(vinylpyrrolidone) (PVP). The diameters of copper NPs prepared by chemical reduction procedure are in the range of 5-10 nm, and copper NPs disperse well with no clear aggregation .
Surface modification is always used to enhance the dispersibility of NPs in the preparation of nanofluids. For example, diamond NPs (DNPs) were purified and surface modified by acid mixtures of perchloric acid, nitric acid and hydrochloric acid according to the literature  before being dispersed into the base fluids. SiC NPs were heated in air to remove the excess free carbon and their surfaces modified to enhance their dispersibility.
Consideration on the thermal conductivity measurement
Inconsistent experimental results and controversial arguments arise unceasingly from different groups conducting research on nanofluids, indicating the complexity of the thermal transport in nanofluids. Through an investigation, a large degree of randomness and scatter have been observed in the experimental data published in the open literature. Given the inconsistency in the data, it is impossible to develop a convincing and comprehensive physical-based model that can predict all the trends. To clarify the suspicion on the scattered and wide-ranging experimental results of the thermal conductivity obtained by different groups, it is preferred to screen the measurement technique and procedure to guarantee the accuracy of the obtained results.
Several researchers observed the “time-dependent characteristic” of thermal conductivity [34-36], that is to say, thermal conductivity was the highest right after nanofluid preparation, and then it decreased considerably with elapsed time. We believe that the “time-dependent characteristic” does not represent the essence of thermal conduction capability of nanofluids. The following two factors may account for this phenomenon. The first one is the motion of the remained particle caused by the agitation during the nanofluid preparation. To make a nanofluid homogeneous and long-term stable, it is always subjected to intensive agitation including magnetic stirring and sonication to destroy the aggregation of the suspended NPs. In very short time after nanofluid preparation, the NPs still keep moving in the base fluid (different from Brownian motion). The motion of the remained particle would cause convection and enhance the energy transport in the nanofluids. Second, when a nanofluid is subjected to long-time sonication, its temperature would be increased. The temperature goes down gradually to the surrounding temperature (thermal conductivity measurement temperature). In very short time after the sonication stops, the process has been remaining. Although the temperature decrease is not severe, the thermal conductivity obtained is very sensitive to the temperature decrease when the transient hot-wire technique is used to measured the thermal conductivity. In our measurements, this phenomenon would be observed. When measuring the thermal conductivity at an unequilibrium state, it was found that the measured data might be very different for a nanofluid even at a specific temperature (see 25°C) if the process to reach this temperature is different. If the temperature is increasing, then the datum obtained of the thermal conductivity would be lower than the true value. While the temperature is decreasing, the datum obtained of the thermal conductivity would be higher than the true value. Therefore, keeping a nanofluid stable and initial equilibrium is very important to obtain accurate thermal conductivity data in measurements.
A transient short hot-wire method was used to measure the thermal conductivities of the base fluids (k0) and the nanofluids (k). The detailed measurement principle, procedure, and error analysis have been described in . In our measurements, a platinum wire with a diameter of 50 μm was used for the hot wire, and it served both as a heating unit and as an electrical resistance thermometer. The platinum wire was coated with an insulation layer of 7-μm thickness. Initially the platinum wire immersed in media was kept at equilibrium with the surroundings. When a regulation voltage was supplied to initiate the measurement, the electrical resistance of the wire changed proportionally with the rise in temperature. The thermal conductivity was calculated from the slope of the rise in the wire’s temperature against the logarithmic time interval. The uncertainty of this measurement is estimated to be within ± 1.0%. A temperature-controlled bath was used to maintain different temperatures of the nanofluids. Instead of monitoring the temperature of the bath, a thermocouple was positioned inside the sample to monitor the temperature on the spot. When the temperature of the sample reached a steady value, the authors waited for further 20 min to make sure that the initial state is at equilibrium. At every tested temperature, measurements were made three times and the average values were taken as the final results. A 20-min interval was needed between two successive measurements. After the above-mentioned careful check on the measurement condition and procedure, the authors could gain confidence on the experimental results.
Influencing factors of thermal conductivity enhancement
In the experiment of the study, it was found that the thermal conductivity enhancements of nanofluids might be influenced by multi-faceted factors including the volume fraction of the dispersed NPs, the tested temperature, the thermal conductivity of the base fluid, the size of the dispersed NPs, the pretreatment process, and the additives of the fluids. The effects of these factors are presented in this section.
The idea of nanofluid application originated from the fact that the thermal conductivity of a solid is much higher than that of a liquid. For example, the thermal conductivity of the most used conventional heat transfer fluid, water, is about 0.6 W/m · K at room temperature, while that of copper is higher than 400 W/m · K. Therefore, particle loading would be the chief factor that influences the thermal transport in nanofluids. As expected, the thermal conductivities of the nanofluids have been increased over that of the base fluid with the addition of a small amount of NPs. Figure 1 shows the enhanced thermal conductivity ratios of the nanofluids with NPs at different volume fractions [7,8,38-42]. (k - k0)/k0 and φ refer to the thermal conductivity enhancement ratio of nanofluids and the volume fraction of NPs, respectively, in this article. Figure1a presents oxide nanofluids, while Figure 1b presents nonoxide nanofluids. The results show that all the nanofluids have noticeable higher thermal conductivities than the base fluid without NPs. In general, the thermal conductivity enhancement increases monotonously with the volume fraction. For the graphene nanofluid with a volume fraction of 0.05, the thermal conductivity can be enhanced by more than 60.0%. There is an approximate linear relationship between the thermal conductivity enhancement ratios and the volume fraction of graphene nanosheets. The nanofluids containing graphene nanosheets show larger thermal conductivity enhancement than those containing oxide NPs. It demonstrates that graphene nanosheet is a good additive to enhance the thermal conductivity of base fluid. However, the enhancement ratios of nanofluids containing graphene nanosheets are less than those of CNTs with the same loading. Many factors have direct influence on the thermal conductivity of the nanofluid. One of the important factors is the crystal structure of the inclusion in the nanofluid. Graphene is a one-atom-thick planar sheet of sp2-bonded carbon atoms that are densely packed in a honeycomb crystal lattice. The perfect structure of graphene is damaged when graphite is chemically oxidized by treatment with strong oxidants. There is no doubt that the high thermal conductivity is diminished by defects, and the defects have direct influence on the heat transport along the 2-D structure.
Figure 1. Thermal conductivity enhancement ratios of the nanofluids as a function of nanoparticle loading. (a) Oxide nanofluids: MgO-EG ; Al2O3-EG ; ZnO-EG ; (b) Nonoxide nanofluids: CNT-EG ; DNP-EG ; Graphene-EG ; Cu-EG .
Some studies have demonstrated that the temperature has a great effect on the enhancement of the thermal conductivity for nanofluids. However, there is considerable disagreement in the literature with respect to the temperature dependence of their thermal conductivity. For example, Das et al. reported strong temperature-depended thermal conductivity for water-based Al2O3 and CuO nanofluids . The thermal conductivity enhancements of nanofluids containing Bi2Te3nanorods in FC72 and in oil had been experimentally found to decrease with increasing temperature . Micael et al. measured the thermal conductivities of EG-based Al2O3 nanofluids at temperatures ranging from 298 to 411 K. A maximum in the thermal conductivity was observed at all mass fractions of NPs .
Figure 2 shows our measured temperature-depended thermal conductivity enhancements of nanofluids [8,38-42]. For EG-based nanofluids containing MgO, ZnO, SiO2, and graphene NPs, the thermal conductivity enhancements almost remain constant when the tested temperature changes (see Figure 2a), which means that the thermal conductivity of the nanofluid tracks the thermal conductivities of the base liquid in the experimented temperature range of this study. The thermal conductivity enhancements of DW-EG-based nanofluids containing MgO, ZnO, SiO2, Al2O3, Fe2O3, TiO2, and graphene NPs also appear to have the same behavior. It was further found that kerosene-based Fe3O4 nanofluids presented temperature-independent thermal conductivity enhancements. Patel et al.  reported that the thermal conductivity enhancement ratios of Cu nanofluids are enhanced considerably when the temperature increases. The experimental results of this study shown in Figure 2b demonstrated similar tendency. At 10°C, the thermal conductivity enhancement of EG based Cu nanofluid with 0.5% nanoparticle loading is less than 15.0%. When the temperature is increased to 60°C, the enhancement reaches as large as 46.0%. Brownian motion of the NPs has been proposed as the dominant factor for this phenomenon. For the EG-based CNT nanofluids, cylindrical nanotubes with large aspect ratios were used as additions. The effect of Brownian motion will be negligible. Typical conduction-based models will give (k - k0)/k0, independent of the temperature. However, results shown in Figure 2b illustrate that (k - k0)/k0increases, though not drastically, with the temperature. CNT aggregation kinetics may contribute to the observed differences . It is worthy of bearing in mind that the temperatures of the base fluid and the nanofluid should be the same when compared with the thermal conductivities between them. Comparison of the thermal conductivities between the nanofluid at one temperature and the base at another one is meaningless.
Figure 2. Thermal conductivity enhancement varying with the tested temperatures. (a) Oxide nanofluids: MgO-EG ; ZnO-EG; Graphene-EG ; (b) Nonoxide nanofluids: Cu-EG ; CNT-EG; DNP-EG .
Figure 3 shows the relation between the enhanced thermal conductivity ratios of the nanofluids and the thermal conductivities of the base fluids [7,8,40,41]. It is clearly seen that no matter what kind of nanoparticle was used, the thermal conductivity enhancement decreases with an increase in the thermal conductivity of the base fluid. For pump oil (PO)-based Al2O3 nanofluid with 5.0% nanoparticle loading, the thermal conductivity can be enhanced by more than 38% compared to that of PO. When the base fluid is substituted with water, the thermal conductivity enhancement achieved is only about 22.0% . A greater dramatic improvement in thermal conductivity of CNT nanofluid is seen for a base fluid with lower thermal conductivity. At 1.0% nanoparticle loading, the thermal conductivity enhancements are 19.6, 12.7, and 7.0% for CNT nanofluids in decene, EG, and DW, respectively. No matter what kind of base fluid is used, the thermal conductivity enhancement of CNT nanofluids is much higher than that for Al2O3 nanoparticle suspensions  at the same volume fraction. The reason would lie in the substantial difference in thermal conductivity and morphology between alumina nanoparticle and carbon nanotube.
Figure 3. Thermal conductivity enhancement ratios as a function of the thermal conductivities of the base fluids: Al2O3 NFs ; CNT NFs ; Graphene NFs ; DNP NFs .
Figure 4 presents the thermal conductivity enhancement of the nanofluids as a function of the specific surface area (SSA) of the suspended particles . It is seen that the thermal conductivity enhancement increases first, and then decreases with an increase in the SSA, with the largest thermal conductivity at a particle SSA of 25 m2 · g-1. We ascribe the thermal conductivity change behavior to twofold factors. First, as particle size decreases, the SSA of the particle increases proportionally. Heat transfer between the particle and the fluid takes place at the particle-fluid interface. Therefore, a dramatic enhancement in thermal conductivity is expected because a reduction in particle size can result in large interfacial area. Second, the mean free path in polycrystalline Al2O3 is estimated to be around 35 nm, which is comparable to the size of the particle that was used. The intrinsic thermal conductivity of nanosized Al2O3 particle may be reduced compared to that of bulk Al2O3 due to the scattering of the primary carriers of energy (phonon) at the particle boundary. It is expected that the suspension’s thermal conductivity is reduced with an increase in the SSA. Therefore, for a suspension containing NPs at a particle size much different from the mean free path, the thermal conductivity increases when the particle size decreases because the first factor is dominant. However, when the size of the dispersed NPs is close to or smaller than the mean free path, the second factor will govern the mechanism of the thermal conductivity behavior of the suspension.
Figure 4. Enhanced thermal conductivity ratios as a function of the SSAs: Al2O3-EG ; Al2O3-PO .
Figure 5 depicts the thermal conductivity enhancements of nanofluids containing CNTs with different sizes . The base fluid is DW, and the volume fraction of the CNTs is 0.0054. It is observed from Figure 5 that the thermal conductivity enhancements show differences among these three kinds of nanofluids containing SWNTs, DWNTs, and MWNTs as the volume fraction of CNTs is the same. Two influencing factors may be addressed. The first one is the intrinsic heat transfer performance of the CNTs. It is reported that the thermal conductivity of CNTs decreases with an increase in the number of the nanotube layer. The tendency of the thermal conductivity enhancement of the obtained CNT nanofluids accords with that of the heat transfer performance of the three kinds of CNTs. The second one is the alignment of the liquid molecules on the surface of CNTs. There are greater number of water molecules close to the surfaces of CNTs with smaller diameter due to the larger SSA if the volume fractions of CNTs are the same. These water molecules can form an interfacial layer structure on the CNT surfaces, increasing the thermal conductivity of the nanofluid .
Figure 5. Thermal conductivity enhancements of nanofluids containing CNTs with different sizes: SWNT-DW ; DWNT-DW; MWNT-DW .
In the preparation of nanofluids, solid additives are always subjected to various pretreatment procedures. The initial incentive is to tailor the surfaces of the NPs to enhance their dispersibility, thereby to enhance the stability of the nanofluids. The morphologies would be significantly changed when CNTs were subjected to chemical or mechanical treatments. Theoretical research into the thermal conductivity of composites containing cylindrical inclusions has demonstrated that the morphologies, including the aspect ratio, have influence on the effective thermal conductivity of the composites. Therefore, it can be expected that the thermal conductivity of CNT contained nanofluids would be affected by the pretreatment process.
Figure 6 shows the dependence of the thermal conductivity enhancement on the ball milling time of CNTs suspended in the nanofluids . From theoretical prediction, the thermal conductivity of a composite increases with the aspect ratio of the included solid particles [49-51]. Intuition suggests that increasing the milling time should therefore decrease (k - k0)/k0 because of the reduced aspect ratio. Figure 6, however, shows clear peak and valley values in the thermal conductivity enhancement with respect to the milling time for all the studied CNT loadings. For nanofluid at a volume fraction of 0.01, the thermal conductivity enhancements present a peak value of 27.5% and a valley value of 10.4% when the milling times are 10 and 28 h, respectively. The maximal enhancement is intriguingly more than two and half times as the minimal one. Interestingly, when further increased the milling time from 28 to 38 h, (k - k0)/k0 increases from the valley value of 10.4 to 12.8%. Though the increment is not pronounced, it illustrates a difference in tendency from that in the milling time range from 10 to 28 h. Temperature-dependent thermal conductivity enhancement data further indicate that, at all the measured temperatures, nanofluid with CNTs milled for 10 h has the largest increment in thermal conductivity. Glory et al.  reported that the enhancement of the thermal conductivity noticeably increases when the nanotube aspect ratio increases. However, the thermal conductivity enhancement behavior of our CNT nanofluid is very different and cannot be explained only by the effect of the aspect ratio.
Figure 6. Dependence of the thermal conductivity enhancement on the ball milling time of CNTs suspended in the nanofluids .
The above results suggest other dominant factors that have the influence over the thermal conductivity of the CNT nanofluids. The authors proposed that the nonstraightness and the aggregation would play significantly roles. As is known, the walls of CNTs have similar structure of graphene sheet, and the thermal conductivity of CNTs shows greatly anisotropic behavior. Heat transports substantially quicker through axial direction than through radial direction . For a nonstraight CNT, the high thermal anisotropy of CNTs induces a unique property that individual CNTs are nearly perfect one-dimensional thermal passages with negligibly small heat flux losses during long distance heat conductions . For a nonstraight CNT with length L under a two-end temperature difference, the heat flux q goes through a curled passage. This CNT can be regarded as an equivalent straight thermal passage with a distance of Le. The same heat flux q is conducted between the two ends of this straight passage. Obviously, the equivalent length Le depends on the curvature of the actual nanotube in the nanofluid. A concept, straightness ratio η (η = Le/L), can be adopted to describe the straightness of a curled CNT. The lowest straightness ratio arises when a suspended nanotube forms ring closure .
When subjected to ball milling, CNTs were broken and cut short with appropriate average length. The straightness ratio was significantly increased and heat transports more effectively through the CNTs and across the interfaces between the CNT tips and the base fluid, resulting in the highest thermal conductivity enhancement in a nanofluid containing CNTs milled for 10 h. For nanofluids containing relatively straight nanotubes, the influence of the aspect ratio will surpass that of straightness ratio. Therefore, by further treatment on nanotubes with relatively high straightness ratio, the excessive deterioration of the aspect ratio would decrease the thermal conductivity of nanofluids, causing (k - k0)/k0 decrease from 10 to 28 h. Recent theoretical analysis has revealed that the aggregation of nanoparticle plays a significant role in deciding (k - k0)/k0 . Percolation effects in the aggregates, as highly conducting nanotubes touch each other in the aggregate, help in increasing the thermal conductivity. Our experiments demonstrate that aggregates are the dominant appearance of CNTs when the ball-milling time is increased to 38 h. The aggregation accounts for the increment of thermal conductivity enhancement when the ball-milling time is increased from 28 to 38 h. This result implies that the positive influence of the aggregation surpasses the negative influence of the aspect ratio deterioration.
For some nanofluids, the pH values of the suspensions have direct effects on the thermal conductivity enhancement. Figure 7 presents the thermal conductivity enhancement ratios at different pH values [7,40]. The results show that the enhanced thermal conductivity increases with an increase in the difference between the pH value of aqueous suspension and the isoelectric point of Al2O3 particle . When the NPs are dispersed into a base fluid, the overall behavior of the particle-fluid interaction depends on the properties of the particle surface. For Al2O3 particles, the isoelectric point (pHiep) is determined to be 9.2, i.e., the repulsive forces among Al2O3 particles is zero, and Al2O3 particles will coagulate together under this pH value. Therefore, when pH value is equal or close to 9.2, Al2O3 particle suspension is unstable according to DLVO theory . The hydration forces among particles increase with the increasing difference of the pH value of a suspension from the pHiep, which results in the enhanced mobility of NPs in the suspension. The microscopic motions of the particles cause micro-convection that enhances the heat transport process. Wensel’s study showed that the thermal conductivity of nanofluids containing oxide NPs and CNTs with very low percentage loading decreased when the pH value is shifted from 7 to 11.45 under the influence of a strong outside magnetic field .
Figure 7. Thermal conductivity enhancement ratios at different pH values: Al2O3-DW ; DNP-EG .
For DNP-EG nanofluids, it is observed from Figure 7 that the thermal conductivity enhancement increases with pH values in the range of 7.0-8.0. When pH value is above 8.0, there is no obvious relationship between pH value and the thermal conductivity enhancement. In our opinion, the influence of pH value on thermal conductivity is that pH value has a direct effect on the stability of nanofluids. When pH value is below 8.5, the suspension is not very stable, and DNPs are easy to form aggregations. The alkalinity of the solution is helpful to the dispersion and the stability of the nanofluids. In order to verify the above statement, the influence of settlement time on the thermal conductivity enhancement was further investigated. It is found that the thermal conductivity enhancement decreases with elapsed time for DNP-EG nanofluid when pH is 7.0. However, for the stable DNP-EG nanofluids with pH of 8.5, there is no obvious thermal conductivity decrease for 6 months .
Surfactant addition is an effective way to enhance the stability of nanofluids. Kim’s study revealed that the thermal conductivity decreased rapidly for the instable nanofluids without surfactants after preparation. However, no obvious changes in the thermal conductivity of the nanofluids with sodium dodecyl sulfate (SDS) as surfactant were observed even after 5-h settlement . Assael et al. investigated the thermal conductivities of the aqueous suspension of CNTs. When Sodium dodecyl sulfate (SDS) was employed as the dispersant, the maximum thermal conductivity enhancement obtained was 38.0% for a nanofluid with 0.6 vol% CNT loadings . When the surfactant is substituted with hexadecyltrimethyl ammonium bromide (CTAB), the maximum thermal conductivity enhancement obtained was 34.0% for same fraction of CNT loading . Liu et al. reported that the thermal conductivity of carbon nanotube-synthetic engine oil suspensions is higher compared with that of same suspensions without the addition of surfactant. The presence of surfactant as stabilizer has positive effect on the carbon nanotube-synthetic engine oil suspensions.
We used cationic gemini surfactants (12-3(4,6)-12,2Br-1) to stabilize water-based MWNT nanofluids. These surfactants were prepared following the process described in . Figure 8presents the thermal conductivity enhancement ratios of the CNT-contained nanofluids with different surfactant concentrations. The volume fraction of the dispersed CNTs is 0.1%. The critical micelle concentration of 12-3-12, 2Br-1 is reported as 9.6 ± 0.3 × 10-4 mol/l . Ten times critical micelle concentration of 12-3-12, 2Br-1 is 0.6 wt%. Solutions of 12-3-12, 2Br-1 with different concentrations (0.6, 1.8, and 3.6 wt% at room temperature) were selected to prepare CNT nanofluids. It is observed that at all the measured temperatures the thermal conductivity enhancement decreases with the surfactant addition. The surfactant added in the nanofluids acts as stabilizer which improves the stability of the CNT nanofluids. However, excess surfactant addition might hinder the improvement of the thermal conductivity enhancement of the nanofluids.
Figure 8. Thermal conductivity enhancement ratios with different surfactant concentrations.
The effect of the structures of cationic gemini surfactant molecules on the thermal conductivity enhancement is shown in Figure 9. The fractions of the dispersed CNTs and the cationic gemini surfactants is 0.1 vol% and 0.6 wt%, respectively. The spacer chain length of the cationic gemini surfactant increase from 3 methylenes to 6 methylenes. It is seen that the thermal conductivity enhancement ratio increases with the decrease of spacer chain length of cationic gemini surfactant. Zeta potential analysis indicates that the CNT nanofluids stabilized by gemini surfactant with short spacer chain length have better stabilities. Increase of spacer chain length of surfactant might give rise to sediments of CNTs in the nanofluids, resulting in the decrease of thermal conductivity enhancement of the nanofluids.
Figure 9. Effect of surfactant structures on the thermal conductivity enhancement ratio.
Nanofluids have great potential for heat transfer enhancement and are highly suited to application in practical heat transfer processes. This provides promising ways for engineers to develop highly compact and effective heat transfer equipments. More and more researchers have paid their attention to this exciting field. When addressing the thermal conductivity of nanofluids, it is foremost important to guarantee the accuracy in the measurement of the thermal conductivity of nanofluids. Two aspects should be considered. The first one is to prepare homogeneous and long-term stable nanofluids. The second one is to keep the initial equilibrium before measuring the thermal conductivity. In general, the thermal conductivity enhancement increases monotonously with the particle loading. The effect of temperature on the thermal conductivity enhancement ratio is somewhat different for different nanofluids. It is very important to note that the temperatures of the base fluid and the nanofluid should be the same while comparing the thermal conductivities between them. With an increase in the thermal conductivity of the base fluid, the thermal conductivity enhancement ratio decreases. Considering the effect of the size of the inclusion, there exists an optimal value for alumina nanofluids, while for the CNT nanofluid, the thermal conductivity increases with a decrease of the average diameter of the included CNTs. The thermal characteristics of nanofluids might be manipulated by means of controlling the morphology of the inclusions, which also provide a promising way to conduct investigation on the mechanism of heat transfer in nanofluids. The additives like acid, base, or surfactant play considerable roles on the thermal conductivity enhancement of nanofluids.
CNTs: carbon nanotubes; DNPs: diamond NPs; DW: deionized water; DWNTs: double-walled CNTs; EG: ethylene glycol; KOH: potassium hydroxide; LP: liquid paraffin; MWNTs: multi-walled CNTs; NPs: nanoparticles; PVP: poly(vinylpyrrolidone); SDS: sodium dodecyl sulfate; SHW: short hot wire; SSA: specific surface area; SWNTs: single-walled CNTs; THW: transient hot wire; TCNTs: treated CNTs.
The authors declare that they have no competing interests.
HQ supervised and participated all the studies. He wrote this paper. WY carried out the studies on the nanofluids containing copper nanoparticles, graphene, diamond nanoparticles, and several kinds of oxide nanoparticles. YL carried out the studies on the nanofluids containing other oxide nanoparticles. LF carried out the studies on the nanofluids containing carbon nanotubes.
This study was supported by the National Science Foundation of China (50876058), Program for New Century Excellent Talents in University (NCET-10-883), and the Program for Professor of Special Appointment (Eastern Scholar) at Shanghai Institutions of Higher Learning.
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Huaqing Xie*, Wei Yu, Yang Li and Lifei Chen
School of Urban Development and Environmental Engineering, Shanghai Second Polytechnic University, Shanghai 201209, China
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Nanoscale Research Letters 2011, 6:124 doi:10.1186/1556-276X-6-124
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Tweet Microporous Silica Based Membranes for Desalination Abstract: This review provides a global overview of microporous silica based membranes for desalination via pervaporation with a focus on membrane synthesis and processing, transport mechanisms and current state of the art membrane performance. Most importantly, the recent development and novel concepts for improving the hydro-stability and separating [...]
Microporous Silica Based Membranes for Desalination
Abstract: This review provides a global overview of microporous silica based membranes for desalination via pervaporation with a focus on membrane synthesis and processing, transport mechanisms and current state of the art membrane performance. Most importantly, the recent development and novel concepts for improving the hydro-stability and separating performance of silica membranes for desalination are critically examined. Research into silica based membranes for desalination has focussed on three primary methods for improving the hydro-stability. These include incorporating carbon templates into the microporous silica both as surfactants and hybrid organic-inorganic structures and incorporation of metal oxide nanoparticles into the silica matrix. The literature examined identified that only metal oxide silica membranes have demonstrated high salt rejections under a variety of feed concentrations, reasonable fluxes and unaltered performance over long-term operation. As this is an embryonic field of research several target areas for researchers were discussed including further improvement of the membrane materials, but also regarding the necessity of integrating waste or solar heat sources into the final process design to ensure cost competitiveness with conventional reverse osmosis processes.
Keywords: desalination; pervaporation; microporous silica; metal oxide silica; hybrid silica; carbon template silica
Water is essential for life and the rapid increase in the global population, and corresponding urbanization has seen the demand for both the quantity and quality of fresh water increase dramatically. One of the major challenges of the 21st century, if not the most important of all, is water scarcity, with the security of social and economic development of a country closed linked to its water resources. Nearly every industrial sector is dependent upon the availability of water, and water shortages have a resounding impact on all levels of society from the general public to health and politics. Indeed, the major problems encountered by water shortages include drought and famine, loss of production in primary industries, loss of job opportunities, poor health and hygiene as well as an increase in the cost of fresh water. This situation is made more complex by the fact that, according to the World Health Organization, more than 15% of the world’s population have no access to potable water and more than 37% have no access to sanitation . Against this backdrop, desalination is becoming an increasingly important tool in the fight to the global demand for clean water.
Membrane technologies have long been an attractive approach to separation in industry, because they are fast and relatively energy efficient processes. In addition, they frequently offer high operational stability, low operating costs and are simple to integrate and control within larger industrial process trains. Indeed, they have been successfully applied to the desalination industry with such vigor that they have long overtaken traditional thermal processes to become the gold standard . In general, there are three main types of membrane processes that are currently applied including reverse osmosis (RO), membrane distillation (MD) and pervaporation (PV) . RO depends on the ability of the ‘dense’ membrane to repel salt ions whilst allowing the passage of water molecules. The transport is governed by a solution-diffusion mechanism with the driving force being an external pressure difference large enough to overcome the osmotic pressure of the salt water. On the other hand, MD is a thermal process that requires a porous, hydrophobic membrane wherein the passage of water vapour only is permissible. PV, by contrast, uses molecular sieve type of membranes that allows only passage to water molecules but relies on a water vapour pressure difference. Both of these desalination processes require very different types of membranes with vastly different properties and configurations. Currently, there are two main types of membranes for water desalination, namely polymeric (e.g., polyamide-, polysulfone-, polyfurane- and cellulose-based for RO and polytetrafluoroethylene for MD) and inorganic composite or ceramic membranes (alumina-, zirconia-, titania-, zeolite-, silica- and carbon-based). Between these two classes of membranes, polymeric membranes are the most mature and well-established in the desalination industry due to their low cost, manufacturability, simple module design and improved permeability and selectivity [4,5]. However, these membranes suffer from swelling phenomenon, a short life-span due to biofouling as well as poor thermal and chemical resistance .
Inorganic membranes, on the other hand, are more resistant to process conditions. In addition, they are by their very nature, porous and hence desalinate via different transport mechanisms to polymeric membranes, based primarily on their pore size. In particular, zeolites and amorphous silica based membranes are attractive candidates for water desalination due to the advantages of their tunable pore sizes and morphology thereby offering higher selectivity. Furthermore, interest in amorphous silica based membranes is gaining momentum because of their simple fabrication techniques, relatively low cost and excellent molecular sieving properties as demonstrated in studies where they are utilized to separate gas molecules [6–9]. In these cases, microporous silica membranes have molecular-sieving structures with pore sizes on the order of the kinetic diameter of the species to be separated (dp = 3–5 Å) and therefore the membrane acts via PV as selective barrier between the water molecule (dk = 2.6 Å) and the hydrated salt ions (e.g., Na+: dk = 7.2 Å and Cl−: dk = 6.6 Å) [10,11], thus allowing the separation of water and salt. However, due to the amorphous nature of the silica material, when exposed to water the silica matrix may undergo dissolution and/or densification . This is a major problem for using silica based membranes in desalination as the effect decreases the overall separation performance and ultimately the quality of the desalinated water. Therefore, a concerted effort has been devoted to improving the hydro-stability of these membranes for various industrial applications.
Many recent reviews have been published for membrane desalination and desalination technologies which are both exhaustive and comprehensive [2,4,5,13–16]. Amongst them, polymeric membranes and zeolites have played a major role. Thus, the contribution of this review is to cover recent studies of non-crystalline microporous silica based membranes for desalination and the new strategies focusing on improving hydro-stability and membrane properties for potential water desalination applications.
2. Membrane Processing and Transport Mechanisms for Water Desalination
Water desalination is a process in which fresh water is extracted from aqueous solutions such as seawater, brackish water and brine, which contain dissolved salts and other minerals. For water molecules to diffuse through a membrane, a driving force must be established, otherwise water molecules will remain mixed in the aqueous salt solution. The driving force is associated with concentration, pressure and temperature difference between the feed and permeate sides of the membrane. In the case of RO processes, the water molecules must overcome the osmotic pressure to diffuse through dense polymeric membranes. As the osmotic pressure of typical saline solutions ranges from 0.2 MPa to 3 MPa for brackish water to seawater respectively, RO desalination processes are generally pressure intensive with pressures of between 6 MPa and 8 MPa commonly used for seawater applications . In contrast, MD does not attempt to overcome the osmotic pressure and so does not require a pressurised feed, although being a thermal process the water flux is proportional to the vapour pressure difference across the membrane. MD generally uses porous hydrophobic membranes, where pore size ranges between 1 µm and 100 Å, and the water vapour permeating via the pores is subsequently condensed downstream to produce fresh water . MD operates at lower temperatures (up to 70 °C) when compared to conventional thermal process such as multi-stage flash or multi-effect distillation. Finally, the PV process, when applied to desalination, employs molecular-sieving (dp = 3–5 Å) ceramic membranes with a narrow pore distribution smaller than the diameter of the hydrated salt ions (>6 Å). Therefore they have the potential to completely reject salt ions while permitting water molecules to permeate. MD and PV are similar processes that can be chiefly identified by the way in which the membrane functions. If the membrane is simply a support structure that allows a meniscus to form on the feed side and plays no role in separation then the process is MD. If however, the membrane actively participates in the separation process then the process is PV. To further provide clarity between these three membrane processes, Figure 1 shows a diagram as comparison of RO, MD and PV in desalination processes.
Figure 1. Schematic representation of transport mechanism through a membrane via
(A) reverse osmosis, (B) membrane distillation and (C) pervaporation for seawater desalination .
PV is a well-established water separation technique particularly in alcohol dehydration, although under those circumstances dense polymeric membranes are typically employed . In PV separation processes, the transport resistance is governed by the sorption equilibrium and mobility of water molecules in the silica membrane based on a molecular sieving mechanism [3,15,16,19,20]. Therefore, the transport of the larger hydrated salt ions is excluded through the membrane . In a typical PV process, the membrane acts as a molecular scale selective barrier between the two phases which consist of the liquid phase in the feed and the vapour phase in the permeate side. In order to create a driving force, vacuum is applied on the permeate side of the membrane while the feed side is kept at atmospheric pressure and temperature. The water molecules permeate through the membrane to the exclusion of the salt ions, evaporate on the permeate side and are then convectively transported to the condenser. Fundamentally, the condenser functions to reduce the water vapour pressure on the permeate side by changing the water phase from vapour to liquid. This function allows for a steady state driving force to be maintained throughout the PV operation.
Similar to MD, PV can operate in several different arrangements. The most common MD operational arrangements have been well reviewed elsewhere . The most common PV arrangements are shown in Figure 2 to provide context and include: (i) vacuum; (ii) air gap and (iii) sweep flow. PV can operate using any setup that allows a vapour pressure gradient to form but does not allow the permeate to flow back into the feed.
Figure 2. Pervaporation (PV) processes in various operational arrangements.
The PV process variables that are commonly investigated include temperature, pressure, total dissolved solids concentration and the ionic strength of the feed solution. The effect of these variables on water transport through the membrane is measured by two important factors which determine the overall membrane performance: (1) flux of the water and (2) selectivity or rejection of the salt ions. The permeate water is captured in a condenser and the flux (kg m−2 h−1), F, of water during a given period of time is calculated using Equation (1):
where M is the permeate mass (kg), S is the membrane surface area (m2) and t is the testing time (h). The salt rejection (%), R, of the membrane is determined by using Equation (2):
R = (Cf - Cp/ Cf) ×100%
where Cf and Cp are the salt concentrations in the feed and permeate solutions, respectively, measured from solution conductivity. Both of the equations are used prolifically in the literature to provide comparison measure for the overall membrane performance in both MD and PV experiments for water desalination. Based on the theory of MD and PV, the salt rejection should equate to 100% since the salt ions will not vapourise under the typical testing conditions. Instead they will crystallize on the inner surface of the membrane on the permeate side if they also find passage across the membrane. There are several reasons that this could occur, but for silica-based membranes this is primarily the result of imperfections in the top layer as a result of poor membrane preparation or silica disintegration in the aqueous environment. Therefore, several research groups have taken this into account by flushing the permeate salt when determining the overall salt rejection [19,22].
As previously alluded to, amorphous silica membranes present an interesting classification problem for membrane desalination technologies, because despite being porous, the water transport through the membrane cannot be described as a conventional MD. One of the major reasons is that in PV using silica based membranes, the pore sizes are too small to effectively form a meniscus associated with a liquid surface tension as it is the case in MD processes. In this case, the Kelvin equation for the liquid-vapour equilibrium is not applicable, as the pure liquid saturation pressure above a convex liquid surface is essentially the same as the pressure above a flat surface. In other words, the pressure of the water molecules at the pore entrance is possibly the same as in the feed bulk liquid phase (i.e., hydrostatic pressure). Having said that, silica based membranes for PV desalination cannot truly be described as activated transport either, as is the case for these membranes in gas separation . Increasing feed bulk liquid pressure results in almost no water flux changes  as expected because changing the bulk feed pressure has a negligible effect on the vapour pressure of the feed; yet changing the vapour pressure of the feed, by increasing its temperature, delivers water flux improvements. Hence, in this case PV closely complies with Darcy’s law (N = K ΔP°) where the water flux (N) is proportional to the water vapour pressure (ΔP°) and coefficient K, which are in turn temperature dependent. Silica derived membranes are hydrophilic materials and the water transport can be described by a sorption-diffusion mechanism. In the case of silica-based membranes for PV, water molecules must preferentially access the pore entrances of the silica matrix to permeate through the membrane, a surface adsorption process. Hence, the water transport can be summarised in four successive steps, namely, (1) selective surface adsorption from the bulk liquid mixture, (2) selective access of water to the pore entrance at the membrane interface on the feed side, (3) diffusion of water from the feed side to the permeate side, (4) desorption of water into vapour phase at the membrane interface of the permeate side. Therefore, the physico-chemical properties of the silica membranes as well as their interaction with the water molecules are equally influential.
3. Features of Silica Based Membranes for Desalination
3.1. Features of Silica Based Membranes for Desalination
Amorphous silica materials that can be tailored to pore sizes in the range of 3–5 Å are highly suitable for selective membranes in water desalination applications. Several techniques have been widely developed to effectively control the pore size of silica derived membranes, including sol-gel methods [24–31] and chemical vapour deposition (CVD) [32–35]. Although remarkable progress in gas separation applications have been reported using both methods, to date only silica membranes derived via sol-gel processes have been investigated for desalination applications. One of the major reasons is that the sol-gel method is one of the most simple and cost effective routes, which still offers the flexibility to tailor the required porosity. Traditionally, the sol-gel method is a wet chemical process to fabricate metal oxide powders starting from a chemical solution which acts as a precursor for an integrated network (gel). This method is frequently adopted in membrane synthesis or membrane pore modification due to its controllability and homogeneity [24,30,36–38], and it includes various steps such as sol preparation, gel formation, drying and thermal treatment. Many types of silicon alkoxide precursors have been utilized, but the clear majority of research describes work using tetraethoxysilane (TEOS) [39–41]. The sol gel synthesis has been well described in a variety of reference materials , and so briefly it involves the hydrolysis (Equation (3)) and condensation reactions (Equations (4) and (5)) of a metal alkoxides to form a network. In the hydrolysis reaction, the alkoxide groups (OR, where R is an alkyl group, CxH2x+1) are replaced with hydroxyl groups (OH). The silanol groups (Si-OH) are subsequently involved in the condensation reaction producing siloxane bonds (Si-O-Si), alcohols (R-OH) and water. The desired microporous structure of the silica layer is thus partially determined by both the reactivity and the size of the precursors, but also by the appropriate selection of the precursor, water, alcohol and catalyst concentrations.
|≡Si-OR + H2O ↔ ≡Si-OH + ROH
|≡Si-OR + HO-Si≡ ↔ ≡Si-O-Si≡ + ROH
|≡Si-OH + HO-Si≡ ↔ ≡Si-O-Si≡ + H2O
Hydrolysis and condensation reactions are commonly catalysed by the use of a mineral base or acid. In the case of a silicon alkoxide, acidic conditions usually produce sols with fractal-like structures which have been shown to be more favorable for the formation of microporous silica with smaller pore sizes . Indeed, when the fractal dimension of those species is low enough, their interpenetration is not restricted during the gelation stage, which gives rise to the formation of weakly-branched structures with small pores . By contrast, basic conditions will otherwise favour the production of highly branched fractal structures and/or colloidal particles. This leads to the production of networks with larger pore sizes and is generally not used to prepare molecular sieveing silica membranes.
3.2. Membrane Preparation
Silica membranes are ultra-thin films (~250 nm) that are traditionally prepared on top of a support for mechanical strength to form an asymmetric structure (as depicted in Figure 3). The support quality plays a major role in the final morphology of the silica derived films as its homogeneity is fundamental in preparing thin films without defects. To achieve this aim, the substrate must have (i) small pore sizes, (ii) low surface roughness and (iii) low defect or void concentration . Substrates with large pores, voids and rough surfaces tend to induce mechanical stress in the films resulting in micro-cracks or pin-hole defects. In order to overcome support roughness, interlayers with smaller pores sizes are typically employed. According to the literature, only a few combinations of support and interlayers have been explored for silica-based membranes for PV desalination. Indeed, supports prepared from α-Al2O3 powders are currently the substrates of choice due to their high porosity and relatively low cost and high mechanical stability. Mesoporous γ-Al2O3, consisting of much smaller pore sizes of ~4 nm are as used in 2 μm thick intermediate layers, and are able to minimize the defect rate observed . However, γ-Al2O3 exhibits low hydrothermal stability , which is of concern if these materials are to be used in applications containing water vapour. Alternate intermediate layers include silica-zirconia composites developed by Tsuru and co-workers [46,47], which are typically more hydro-stable than γ-Al2O3 layers.
The coating of the substrate (or support) using the sol-gel process can be carried out by dip coating, spin coating and the pendulum method. Due to its flexibility to coat both flat and tubular geometries, in addition to small or large substrates, dip coating has been the preferred process to prepare silica based membranes. Scriven  extensively reviewed the dip coating process and proposed five stages: immersion, start-up, deposition, drainage and evaporation. Upon immersion of a substrate to a silica sol, the sol starts adhering to the surface of the substrate. During the withdrawal step, the sol deposits on the surface of the substrate leading to drainage of excess liquid and evaporation of the sol to forming a gel on the support surface. Brinker et al.  proposed that there is a sequential order of structural development that results from drainage accompanied by solvent evaporation, continued condensation reactions and capillary collapse. According to Brinker et al.  the concentration of the deposited film increases 18–36 fold due to evaporation. This causes the formed film to undergo very fast gelation and drying, thus suggesting structural reorganization of the film matrix.
Figure 3. SEM micrograph of the cross-section of a high quality asymmetric membrane structure—Reproduced by permission of The Royal Society of Chemistry (http://dx.doi.org/10.1039/B924327E) .
The withdrawal speed of the substrate from a sol, in addition to the viscosity of the sol, plays an important role in determining the silica thin film formation. Generally, withdrawal speeds reported by several research groups vary between 1 and 20 cm min−1, whilst prepared sols are diluted with ethanol up to 20 times the original sol volume. In this case (low withdrawal speed and low viscosity), the thickness of a film (h) is proportional to U2/3 (where U is the product of the viscosity and withdrawal speed), in accordance to the Landau and Levich equation . Hence, increasing the speed of withdrawal in the dip coating process will yield thicker films and vice versa. As the production of thicker films tends to lead to cracking upon evaporation and gelation, thinner sols of low viscosity with low withdrawal speeds are preferred.
Upon film coating, the membranes are calcined at high temperatures, generally up to 600 °C, in order to fix the silica structure. Higher temperatures tend to densify the silica film, resulting in extremely low fluxes. The calcination process can lead to thermal stresses between the substrate and the thin silica film, possibly causing film cracking and defects. Hence the heating ramp rate is of considerable importance and is typically low at around 1 °C min−1, although recent developments in rapid thermal processing for silica membranes in other applications are challenging this long held view [53,54]. As the thickness of the silica films are generally in the region of 30–50 nm, and possibly a single film may contain defects caused by either inhomogeneity in the support or interlayers, or calcination stresses, or environmental dust; the dip coating and calcination process is generally repeated at least 2–3 times to produce high quality membranes. As environmental dust affects thin film formation, de Vos and Verweij  demonstrated that the quality of silica membranes was greatly improved by simply coating in a clean room environment.
4. Novel Silica Based Membranes in Desalination
4.1. Hydro-stability and Current Strategies
Owing to the affinity of amorphous silica for water adsorption, silica derived membranes undergo structural degradation when exposed to water, leading to a loss of selectivity . Briefly, silica surface materials are prone to rehydration via a mechanism of physisorption of H2O molecules on silanol groups (Si-OH), followed by reaction with a nearby siloxane (chemisorption) [57,58]. As a result, H2O assists the breakage of siloxane groups, allowing for dissociative chemisorption via the hydrolysis reaction (the reverse of Equation (5)) . Therefore, hydrolysed surface siloxanes may become strained, which act as strong acid–base sites, having a rapid uptake of water and becoming mobile . As the silica seeks to reduce its surface energy under hydrothermal conditions , Duke and co-workers  proposed that the mobile and strained hydrolysed siloxane groups migrate to smaller pores where they undergo re-condensation to block the pore, whilst the larger pores become even larger. Hydro-stability is therefore a serious problem for the deployment of silica based membranes for water desalination. To address this problem, researchers have attempted to modify the surface properties of the silica, to minimize the interaction of water molecules with the membrane structure. A summary of the main strategies employed is displayed in Figure 4.
One strategy to solve this challenging problem is introducing non-covalently bonded, organic templates into the pure silica matrix [62–64]. Indeed, the presence of carbon moieties embedded into the silica framework can prevent the mobility of soluble silica groups under hydrolytic attack and consequently inhibits micropore collapse. This was demonstrated by Duke et al.  who successfully prepared carbonized-template molecular sieve silica membranes (CTMSS) by introducing the ionic surfactant (C6 hexyltriethyl ammonium bromide) during the silica sol synthesis. The carbon moieties trapped in the CTMSS matrix were formed by carbonization of the surfactant under vacuum or an inert atmosphere, leading to a hybrid silica/carbon membrane. Although CTMSS membranes still retained their hydrophilic properties, the resultant membranes showed great potential for attaining hydro-stability without compromising the selectivity for wet gas separation . Based on this approach, CTMSS membranes were subsequently tested for desalination performance, demonstrating high salt rejection from seawater .
In a similar study, Wijaya et al.  investigated the effect of the carbon chain length of ionic surfactants in CTMSS membranes for desalination by preparing sol-gels with hexyltriethyl ammonium bromide (C6), dodecyltrimethyl ammonium bromide (C12) and hexadecyltrimethyl ammonium bromide (C16). It was found that the CTMSS membrane prepared with the surfactant with the longest carbon chain (C16) delivered the highest salt rejection, whilst also given the largest pore volume and surface area, although interestingly, the average pore sizes were similar for the three surfactants used. These results suggest that the embedded carbon has a beneficial role in silica matrices and the amount embedded has a direct impact in terms of desalination performance, since the carbon content of the added surfactant is directly related to the amount of carbon remaining following carbonization. However, if the concentration of ionic surfactants is too high they form micelles  which drastically limits the possibility of using the sol-gel to dip coat substrates. In order to increase the carbon content in the silica framework, Ladewig et al.  proposed the use of a non-ionic surfactant such as a tri-block copolymer like polyethylene glycol–polypropylene glycol–polyethylene glycol (PEG-PPG-PEG), a high molecular weight polymer. Silica samples were mixed with 1–20 wt % PEG-PPG-PEG, and increasing the loading of the tri-block copolymer to 10 wt % effectively doubled the pore volume and surface area compared to pure silica, whilst still maintaining microporosity. Further increases in tri-block copolymer loading to 20 wt % altered the structure of the CTMSS materials to produce mesopores. Of greatest relevance to both the preceding studies and future research directions, the CTMSS membranes prepared with 10 wt % PEG-PPG-PEG (i.e., the highest carbon content sample, whilst still remaining microporous) also delivered high salt rejections and water fluxes.
Figure 4. Schematic representation of various strategies for silica modification.
Another approach to increase the hydrothermal stability of the pure silica membrane is by incorporating terminal methyl groups (≡Si-CH3) via various precursors used during the sol-gel synthesis (Figure 5). This was firstly reported by de Vos et al.  who synthesized methylated silica membranes derived by the copolymerization of TEOS and methyltriethoxysilane (MTES) in the presence of ethanol and water, with an acid catalysis. Again, the membranes were calcined under a non-oxidising environment to retain the carbon moieties in the silica matrix. These membranes showed remarkable stability for alcohol dehydration for 18 months, though severe degradation occurred at testing temperatures of ≥95 °C thereafter . Although the addition of methyl ligand groups to silica rendered hydrophobicity, the counter effect was the formation of larger micropores. Duke et al.  investigated the effect of both methyl ligand and non-ligand C6 surfactant as templates in silica membranes for desalination. They found that the CTMSS membrane outperformed the methylated-silica membrane, suggesting that carbonizing the C6 surfactants led to the formation of smaller pores than the covalently attached methyl groups.
Figure 5. Precursors used for the preparation of pure (TEOS), methylated (MTES) and hybrids (BTESE) silica membranes.
Following on from the methyl ligand work, significant hydrothermal improvement can be achieved when the siloxane bridges (Si-O-Si) are partially replaced by organic bridges (Si-CH2-CH2-Si) such as BTESE in Figure 5. In this method, alkyl groups (ethylene groups in Figure 5) between Si atoms, which cannot be hydrolyzed, can be used as a “spacer” to control the silica network size while minimizing the hydrophilicity of the silica pore surface. The sol synthesis for such membrane layers was first developed by Castricum et al.  and consisted of a two-step acid hydrolysis of BTESE/MTES mixtures. In this work they showed that the durability of the membrane network for the dehydration of n-butanol by PV was greatly improved by incorporating hydrolytically stable organic groups as integral bridging components into the nanoporous silica. These hybrid organosilica membranes were able to withstand long-term PV operation of up to 2 years at 150 °C. Recently, Tsuru et al. reported the potential of such BTESE membranes in RO and PV desalination processes .
Alternate efforts have focused on modifying the silica structure through the addition of metal oxides [73–77]. Recently Lin et al.  reported for the first time the potential of cobalt oxide silica (CoOxSi) membranes for desalination of waters from brackish to brine concentrations. CoOxSi xerogels were synthesised via the sol-gel method using TEOS, cobalt nitrate hexahydrate and hydrogen peroxide, at a range of pH from 3 to 6. The pH was altered by addition of ammonia during the sol-gel process. Initial hydrothermal exposure (<2 days) at 75 °C of xerogels resulted in the reduction of pore volume and surface area, although subsequent exposure proved that the pore structure of the xerogels was no longer significantly altered. The CoOxSi synthesized at pH 5 was the most resistant to the hydrothermal degradation, remaining stable and delivering high salt rejections for 570 hours of testing at temperatures up to 75 °C and NaCl salt concentrations up to 15 wt %.
4.2. Membrane Performance: Effect of Testing Conditions
A summary of the reported membranes performance in term of water flux and salt rejection is listed in Table 1. It must be stressed that comparing these results gives an indication of the general performance only. One should be aware of that these results are dependent upon several parameters related to testing condition including feed concentration, salt used, feed temperature, feed flow rate, cross-flow velocity, permeate vapour pressure and fouling/scaling tendencies. In addition, these listed membranes may have different geometries (flat or tubular and sizes) and architecture (thickness of top film, number interlayers number, porosity and substrate). As such, all these factors play a role in the final performance of the tested membranes.
a Feed pressurizing up to 7 bar and permeate vacuum pumping; b Permeate vacuum pumping, resulting in a pressure difference ΔP across the membrane less than 1bar; * Sea water.
The majority of membranes listed in Table 1 were tested for feed synthetic solutions containing NaCl dissolved in deionised water with concentrations ranging from 0.3 to 3.5 wt % in order to simulate the typical salt concentration of brackish water (0.3–1 wt %) and sea water (3.5 wt %). CTMSS membranes gave similar water fluxes varying from 1.4 to 6.3 kg m−2 h−1 with high salt rejections greater than 84%, depending on the operating conditions. Hybrid membranes (i.e., those prepared with terminal methyl groups or covalently bound carbon bridges) also gave similar water fluxes and salt rejections. The hybrid membranes prepared with BTESE delivered considerable high water fluxes at 34 kg m−2 h−1 at 90 °C and excellent salt rejection 99.9%. However, these membranes were tested at very low salt concentration (NaCl 0.2 wt %) and high feed temperature and when cooler feed temperatures (30 °C) were used, the water fluxes reduced considerably (one order of magnitude). In the only study of its kind so far, CoOxSi based silica membranes were also investigated for brine
processing conditions where the salt concentrations ranged from 7.5 to 15 wt %. In this case, an increase in salt concentration in the feed from 0.3 to 15 wt % resulted in a decline of the permeate flux from 1.8 to 0.55 kg m−2 h−1 at 75 °C. However, despite the high salt feed concentrations, the salt rejection remained high suggesting they were stable under these harsh testing conditions.
Analyzing the results reported in Table 1, the trends are very clear with increasing temperature yielding increased water flux whilst increasing salt concentration results in decreasing water flux. For instance, at 0.3 wt % salt feed concentration, the water flux increased by 77% (from 0.4 to 1.8 kg m−2 h−1) as the feed temperature was raised from 20 °C to 75 °C. This can be explained through the thermodynamics of the system in that as the temperature increases, so does the water vapour pressure in the feed stream, leading to an increase in the driving force for water permeation across the membrane. Likewise the water vapour pressure decreases as a function of the salt concentration, partially explaining the decreased flux observed under seawater and brine feed concentrations. However, the water vapour pressure change as a function of the salt concentration at constant temperature is not large enough to justify the large reduction of flux as reported by several groups in Table 1. For instance, in the case of carbonized template CTMSS (ionic C6), experiment was conducted at a fixed temperature of 20 °C . Indeed, water flux was reduced by more than half (56%) by increasing the feed concentration from 0.3 wt % to 3.5 wt %. In that case, the change in vapour pressure driving force of an ideal salt solution will change from 2.3 kPa to 2.28 kPa, representing a decrease of 0.08%  far smaller than the decline in flux, thus demonstrating that salt and temperature polarization are also likely occurring. In this case a boundary layer of more concentrated salt forms at the membrane surface due to the permeation of water through the membrane being faster than the diffusion for fresh water from the bulk to the membrane surface. Likewise thermal boundary layers can form through the conduction and convection of sensible heat and the transfer of latent heat through the vapourisation of water through the membrane. This phenomenon, along with temperature polarization, is commonly observed for MD processes. Interestingly, temperature polarization, whereby the heat flow across the membrane from conduction and convection is sufficient to reduce the temperature at the membrane surface in comparison to the bulk feed, is the more commonly reported problem . The fact that salt concentration polarization is strongest suggest that (a) the silica-based membrane is more insulating than typical polymeric MD counterparts, and that (b) the cross flow velocities investigated were not sufficient to disturb or reduce the mass transfer boundary layer.
The purity of the water in the permeate stream is a fundamental parameter in terms of potable water. As the salt rejection is generally a ratio of salinities (Equation (2)), a high salt rejection for a high feed salt stream does not necessarily translating into potable water. According to the World Health Organization portable water should have a factor called total dissolved solids (TDS) < 600 ppm with an upper limit of TDS < 1000 ppm . To assess the performance of the membranes in Table 1 in terms of water quality, the permeate water concentration was calculated as shown in Figure 6. All the membranes listed in Table 1 produce good quality drinking water (TDS < 600 ppm) for slightly saline water conditions (0.3 wt %). However, only the CoOxSi silica base membranes were able to meet the requirement of 600 ppm for seawater and brine feed conditions. For those membranes with TDS in excess of 600 or 1000, a second pass becomes necessary to achieve potable water requirements. As discussed previously the theory of PV operation necessitates that the permeate stream should be free of salt, regardless of the feed conditions.
The observation that the vast majority of silica-based membranes tested under PV desalination conditions do not give pure water in the permeate stream is strong evidence that research focusing on improving the hydro-stability of the silica as well as the integrity of the membrane layer itself should continue to receive high priority. However, only a handful of authors have reported preliminary stability measurements as listed in Table 1. In the longest performance evaluation reported so far, Lin et al. sequentially tested cobalt oxide silica (CoOxSi) membranes with solutions containing salt at 1 wt % (288 h), 3.5 wt % (144 h), 7.5wt % (72 h) and 15 wt % (72 h), leading to a total of 575 hours . Despite a significant variation in water flux observed during the first 120 h, the water flux tended to stabilize after 5 days of measurement. This was attributed to initial textural and/or structural changes in the CoOxSi matrix and was also observed in nitrogen sorption and FTIR analyses. However, this long term testing successfully demonstrated the improved hydro-stability of CoOxSi membranes at several temperature points and feed concentrations. In the only other studies reported thus far, Duke et al. reported stable performance over 5 h of the CTMSS (Ionic 6) membrane ; and Ladewig et al. showed stable performance over 12 h, suggesting the benefit of the carbonized templating method to improve the hydro-stability of amorphous silica membranes .
Figure 6. Comparison of water quality in the permeate stream.
4.3. Future Challenges
Silica based membranes for desalination applications are still at the embryonic stages of research and development. Therefore, this type of membrane requires significant improvements to be able to compete against both alternate membranes and alternate technologies. Indeed, the RO process using polymeric membranes is now a mature technology, having undergone major research, development and deployment in the last 30 years. This developmental advantage implies that RO will continue to dominate the large desalination plants around the world. However, RO cannot process all feed concentrations, in particular the pressure requirements for brine processing are prohibitive and can even destroy the polymeric membranes. Thus silica based membranes (especially metal oxide silica membranes, such as CoOxSi) operating under PV conditions, could have a niche market in the processing of brines or even the processing or drying of mineral salts such as potash or lithium brines.
In order to be able to compete against polymeric RO membranes, the water fluxes of silica based membranes for processing seawater (NaCl 3.5 wt %) must be significantly increased, by an order of magnitude on average. At the moment high water fluxes in excess of 20 kg m−2 h−1 have been demonstrated for BTESE silica membranes only, although only for slightly saline feed concentrations (NaCl 0.2 wt %) and high temperatures at 90 °C. This raises the second major impediment to silica based membrane PV, the issue of temperature, and ultimately energy consumption. PV is a thermal process and raising the temperature of feed translates into higher vapour pressures which should likewise increase water flux and water production. The problem here is that heat must be generated to increase the temperature of the water (and ultimately vapourise it), which together with the energy required to condense the water vapour explains why the PV process uses more energy per liter of water produced than RO processes which use only pump energy to pressurize the saline water feed. If this heat is supplied through conventional means, the cost will be prohibitive. However, there are several options available to reduce the cost of energy by utilising waste heat from industrial sites and thermal power plants, salt gradient solar ponds or solar heat [80–84]. These options may be attractive to deploy PV using silica based membranes.
A vital aspect of any membrane technology is long term operation and stability. At the moment, CoOxSi silica membranes have demonstrated stability up to 575 hours of operation. Similar tests must also be undertaken for CTMSS and hybrid silica based membranes to show proof of concept. To some extent, the CoOxSi silica membranes showed superior performance than MFI zeolites, which may be viewed as a competing membrane technology. In a recent study, Dobrek and co-workers  reported the dissolution of both S-1 and ZSM-5 top layers in MFI zeolite membranes after 560 hours testing in PV desalination. This was attributed to the combined effects of ion exchange and water dissolution mechanisms. The loss of membrane performance due to the quality of the saline waters can therefore cause deterioration of the materials such as in zeolite membranes, or fouling and scaling as is the case of polymeric membranes . Currently, there is no fouling work reported for silica based membranes mainly due to the embryonic nature of the testing which has occurred under laboratory conditions using synthetic salt solutions. Given the scale of the problem for RO membranes, this is a problem that will require substantial research to ensure that silica based membranes can be deployed in an industrial context to process saline waters to potable quality.
Microporous silica based membranes have been shown to provide excellent molecular sieving properties for gas separation applications but their reported use in water treatment processes, such as desalination, have been limited, primarily due to the lack of stability when exposed to water. However, innovative concepts have been developed in the last two decades to realize the potential of silica based membranes for desalination via PV. In particular, research into silica based membrane desalination has focussed on three distinct methods of stabilising the structure including carbon templated silica, hybrid organic-inorganic silica and metal oxide silica. Whilst these methods have all been successfully trialed for desalination via PV, only metal oxide silica membranes have demonstrated significant potential with high salt rejections under all feed concentrations, reasonable fluxes and unaltered performance for over 575 hours of operation. Indeed they were the only membranes capable of producing potable water from highly concentrated brine feed streams. The target areas of research for membrane scientists is therefore on the materials development to further improve water fluxes (in order to compete with RO processes), to stabilize the silica structure to ensure no reductions in long term performance and to produce defect-free membranes to ensure high salt rejections, at low cost. The final challenge for the membrane research community is to establish the conditions under which PV desalination using silica based membranes is most technically and economically viable. The energy requirements of PV systems are considerable in comparison to RO processes and analysis of the thermodynamics indicates that parity will never be reached when utilizing primary energy sources. However, if PV processes are successfully integrated with waste heat or solar heat sources then the technology may be attractive for niche applications such as brine processing or salt recovery. Regardless, the separation and purification of potable water from desalination is a paramount task which the membrane research community must endeavour to address before water supply becomes a global crisis.
Muthia Elma specially thanks for the scholarship provided by the University of Queensland. The authors acknowledge financial support from the Australian Research Council (DP110101185). Simon Smart also acknowledges funding support from the Australian Research Council (DP110103440).
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Muthia Elma, Christelle Yacou, David K. Wang, Simon Smart and João C. Diniz da Costa *
Films and Inorganic Membrane Laboratory, School of Chemical Engineering, The University of Queensland, Brisbane, Queensland 4072, Australia; E-Mails: firstname.lastname@example.org (M.E.); email@example.com (C.Y.); firstname.lastname@example.org (D.K.W.); email@example.com (S.S.)
* Author to whom correspondence should be addressed; E-Mail: firstname.lastname@example.org; Tel.: +62-7-33656960; Fax: +62-7-33654199.
© 2012 by the authors; licensee MDPI, Basel, Switzerland. This article is an open access article distributed under the terms and conditions of the Creative Commons Attribution license (http://creativecommons.org/licenses/by/3.0/).
admin 27 Mar, 2013
TweetSustainability in Water Supply Sustainable water systems should provide adequate water quantity and appropriate water quality for a given need, without compromising the future ability to provide this capacity and quality. Water systems in the realm of sustainable development may not literally include the use of water, but include systems where the use of water [...]
Sustainability in Water Supply
Sustainable water systems should provide adequate water quantity and appropriate water quality for a given need, without compromising the future ability to provide this capacity and quality. Water systems in the realm of sustainable development may not literally include the use of water, but include systems where the use of water has traditionally been required. Examples include waterless toilets and waterless car washes, whose use helps to alleviate water stress and secure a sustainable water supply.
Accessing the sustainability features in water supply, that is to say, the three-fold goals of economic feasibility, social responsibility and environmental integrity, is linked to the purpose of water use. Sometimes, these purposes compete when resources are limited; for example, water needed to meet the demands of an increasingly urban population and those needs of rural agriculture. Water is used (1) for drinking as a survival necessity, (2) in industrial operations (energy production, manufacturing of goods, etc.), (3) domestic applications (cooking, cleaning, bathing, sanitation), and (4) agriculture. Sustainable water supply is a component of integrated water resource management, the practice of bringing together multiple stakeholders with various viewpoints in order to determine how water should best be managed. In order to decide if a water system is sustainable, various economical, social and ecological considerations must be considered.
Surface freshwater is unfortunately limited and unequally distributed in the world. Almost 50% of the world’s lakes are located in Canada alone (UNEP, 2002). In addition, pollution from various activities leads to surface water that is not drinking quality. Therefore, treatment systems (either large scale or at the household level) must be put in place.
Structures such as dams may be used to impound water for consumption. Dams can be used for power generation, water supply, irrigation, flood prevention, water diversion, navigation, etc. If properly designed and constructed, dams can help provide a sustainable water supply. The design should consider peak flood flows (historical and projected for climate change), earthquake faults, soil permeability, slope stability and erosion, silting, wetlands, water table, human impacts, ecological impacts (including wildlife), compensation for resettlement, and other site characteristics. There are various challenges that large-scale dam projects may present to sustainability: negative environmental impacts on wildlife habitats, fish migration, water flow and quality, and socioeconomic impacts resulting from resettled local communities. A sustainability impact assessment should therefore be performed to determine the environmental, economic and social consequences of the construction.
Groundwater accounts for greater than 50% of global freshwater; thus, it is critical for potable water (Lozan et al, 2007). Groundwater can be a sustainable water supply source if the total amount of water entering, leaving, and being stored in the system is conserved. There are three main factors which determine the source and amount of water flowing through a groundwater system: precipitation, location of streams and other surface-water bodies, and evapotranspiration rate; it is thus not possible to generalize a sustainable withdrawal or pumping rate for groundwater (USGS, 1999). Unsustainable groundwater use results in water-level decline, reduced streamflow, and low water quality, jeopardizing the livelihood of effected communities. Various practices of sustainable groundwater supply include changing rates or spatial patterns of ground-water pumpage, increasing recharge to the ground-water system, decreasing discharge from the groundwater system, and changing the volume of groundwater in storage at different time scales (USGS, 1999). A long-term vision is necessary when extracting groundwater since the effects of its development can take years before becoming apparent. It is important to integrate groundwater supply within adequate land planning and sustainable urban drainage systems.
Collecting water from precipitation is one of the most sustainable sources of water supply since it has inherent barriers to the risk of over-exploitation found in surface and groundwater sources, and directly provides drinking water quality. However, rainwater harvesting systems must be properly designed and maintained in order to collect water efficiently, prevent contamination and use sustainable treatment systems in case the water is contaminated. A number of drinking water treatments exist at point-of-use, each with advantages and disadvantages. These include solar treatment, boiling, using filters, chlorination, combined methods such as filtration and chlorination, flocculation and chlorination. Although technically given the Earth’s surface and precipitation, rainwater harvesting can meet global water demand, the solution can most practically be a supplement to sustainable water supply systems given a level of uncertainty (especially with climate change), and competing land-use applications.
Reclaimed water, or water recycled from human use, can also be a sustainable source of water supply. It is an important solution to reduce stress on primary water resources such as surface and groundwater. There are both centralized and decentralized systems which include greywater recycling systems and the use of microporous membranes. Reclaimed water must be treated to provide the appropriate quality for a given application (irrigation, industry use, etc.). It is often most efficient to separate greywater from blackwater, thereby using the two water streams for different uses. Greywater comes from domestic activities such as washing, whereas blackwater contains human waste. The characteristics of the two wastestreams thus differ.
Desalinisation has the potential to provide an adequate water quantity to those regions that are freshwater poor, including small island states. However, the energy demands of reverse osmosis, a widely-used procedure used to remove salt from water, are a challenge to the adaptation of this technology as a sustainable one. The costs of desalination average around 0.81 USD per cubic meter compared to roughly 0.16 USD per cubic meter from other supply sources (USGS, 2010). If desalination can be provided with renewable energies and efficient technologies, the sustainable features of this supply source would increase. Currently, desalination increases operational costs because of the needed energy (and also carbon dioxide emissions); this in turn raises the cost of the final product. In addition, desalination plants can have negative impacts on marine life, and cause water pollution due to the chemicals used to treat water and the discharge of brine.
Bottled water is a 21st century phenomenon whereby mostly private companies provide potable water in a bottle for a cost. In some areas, bottled water is the only reliable source of safe drinking water. However, often in these same locations, the cost is prohibitively expensive for the local population to use in a sustainable manner. Bottled water is not considered an “improved drinking water source” when it is the only potable source available (UN, 2010). When sustainability metrics are used to access bottled water, it falls short in many situations of being a sustainable water supply. Economic costs, pollution associated with its manufacturing (plastic, energy, etc.) and transportation, as well as extra water use, makes bottled water an unsustainable water supply system for many regions and for many brands. It takes 3-4 liters of water to make less than 1 liter of bottled water (Pacific Institute, 2008).
Potable water requires some of the strictest standards of quality in terms of bacteriological and chemical pollutants. These standards are often governed by national governments; international recommendations can be found from the World Health Organization (http://www.who.int/water_sanitation_health/dwq/guidelines/en/index.html). Drinking water must be freshwater and should be free of pathogens and free of harmful chemicals.
Water in Industry
Water is used in just about every industry. Industrial water withdrawls represent 22% of total global water use (significant regional differences). Its use is notable for manufacturing, processing, washing, diluting, cooling, transporting substances, sanitation needs within a facility, incorporating water into a final product, etc. (USGS, 2010). The food, paper, chemicals, refined petroleum, and primary metal industries use large amounts of water (USGS, 2010). A sustainable water supply in industry involves limiting water use through efficient appliances and methods adapted to the particular industry. Rainwater harvesting on-site (including the creation of large pond-like structures), as well as recycling water in industrial processes, can provide a sustainable water supply for industry without straining municipal water supplies. Industry releases organic water pollutants, heavy metals, solvents, toxic sludge, and other wastes into water supply sources. Industry thus has a dual responsibility for internal sustainable water supply and the protection of external water supply sources.
Water in Agriculture
Agriculture uses the largest amount of freshwater on a global scale. It represents roughly 70% of all water withdrawal worldwide, with various regional differences. In the United States, for example, agriculture accounts for over 80% of water consumption (USDA, 2010). The productivity of irrigated land is approximately three times greater than that of rain-fed land (FAO, 2010). Thus, irrigation is an important factor for sustainable agriculture systems. In addition, global food production is expected to increase by 60% from 2000 to 2030, creating a 14% increase in water demand for irrigation (UN, 2005). Agriculture is also responsible for some of the surface and groundwater degradation because of run-off (chemical and erosion-based). It thus has a dual role in sustainable water supply: (1) using water efficiently for irrigation and (2) protecting surface and groundwater supply sources. Techniques for sustainable water supply in agriculture include organic farming practices which limit substances that would contaminate water, efficient water delivery, micro-irrigation systems, adapted water lifting technologies, zero tillage, rainwater harvesting, runoff farming, and drip irrigation (efficient method that allows water to drip slowly to plant roots by using pipes, valves, tubes and emitters).
Domestic Water Uses
The average household needs an estimated 20-50 liters of water per person per day, depending on various assumptions and practices (Gleick, 1996). Reducing water use through waterless toilets, water efficient appliances, and water quantity monitoring, is an important part of sustainability for domestic water supply. Efficient piping systems that are leak-free and well insulated provide a network that is reliable and help to limit water waste. The aforementioned potable water supply sources, with their sustainability features and sustainability challenges, are all relevant to other domestic uses. Since water quality standards are not as strict for household uses as for drinking, there is more flexibility when considering sustainable domestic water supply (including the potential for reclaimed water use).
A water supply system will be sustainable only if it promotes efficiencies in both the supply and the demand sides. Initiatives to meet demand for water supply will be sustainable if they prioritize measures to avoid water waste. Avoiding wastage will contribute to reducing water consumption and, consequently, to delaying the need for new resources.
On the supply side, it is fundamental to enhance operation and maintenance capabilities of water utilities, reducing non-revenue water (NRW), leakages, and energy use, as well as improving the capacity of the workforce to understand and operate the system. It is also necessary to ensure cost-recovery through a fair tariff system and “intelligent” investment planning. In addition, all alternatives to increase the water supply must be analysed considering the entire life cycle.
On the demand side, the adoption of water efficient technology can considerably reduce water consumption. Investments in less water intensive industrial processes and more efficient buildings lead to a more sustainable water supply. Concrete possibilities of economic savings, social benefits (such as the involvement of different sectors of society to reach a common objective, environmental awareness of the population, etc.) and a range of environmental gains make the adoption of water efficient technologies viable.
Sustainable water supply involves a sequence of combined actions and not isolated strategies. It depends on the individual’s willingness to save water, governmental regulations, changes in the building industry, industrial processes reformulation, land occupation, etc. The challenge is to create mechanisms of regulation, incentives and affordability to ensure the sustainability of the system.
Food and Agriculture Organization of the United Nations (FAO). (2010). Water Use in Agriculture. Retrieved from http://www.fao.org/ag/magazine/0511sp2.htm
Gleick, Peter H. (1996). Basic Water Requirements for Human Activities: Meeting Basic Needs.” Water International 21, 2: 83-92.
US Geological Survey. (2010). Industrial Water Use. Retrieved from http://ga.water.usgs.gov/edu/wuin.html
United States Department of Agriculture. (2010). Irrigation and Water Use. Retrieved from http://www.ers.usda.gov/Briefing/WaterUse/
Lozan, Grassl, et al. (2007). The water problem of our Earth: From climate and the water cycle to the human right for water.
UN Water for Life Decade. (2005). United Nations Department of Public Information (32948—DPI/2378—September 2005—10M).
UNEP. (2002). Vital Water Graphics: An Overview of the State of the World’s Fresh and Marine Waters. Retrieved from http://www.unep.org/dewa/assessments/ecosystems/water/vitalwater/.
Pacific Institute. Water Content of Things. The World’s Water 2008-2009.
United Nations (WHO and UNICEF). (2010). Progress on Sanitation and Drinking Water Update 2010. Retrieved from http://www.unicef.org/media/files/JMP-2010Final.pdf.
USGS. (2010). Thirsty? How ’bout a cool, refreshing cup of seawater? Retrieved from http://ga.water.usgs.gov/edu/drinkseawater.html.
USGS. (1999). Sustainability of Ground-Water Resources. Retrieved from http://pubs.usgs.gov/circ/circ1186/pdf/circ1186.pdf.
Waite, Marilyn. (2010). Sustainable Water Resources in the Built Environment. IWA Publishing: London.
Many of the issues in this article are covered in the book, Sustainable Water Resources in the Built Environment, published in 2010, written by Marilyn Waite.
Sustainable Water Resources in the Built Environment covers elements of water engineering and policy making in the sustainable construction of buildings with a focus on case studies from Panama and Kenya. It provides comprehensive information based on case studies, experimental data, interviews, and in-depth research.
The book focuses on the water aspects of sustainable construction in less economically developed environments. It covers the importance of sustainable construction in developing country contexts with particular reference to what is meant by the water and wastewater aspects of sustainable buildings, the layout, climate, and culture of sites, the water quality tests performed and results obtained, the design of rainwater harvesting systems and policy considerations.
The book is a useful resource for practitioners in the field working on the water aspects of sustainable construction (international aid agencies, engineering firms working in developing contexts, intergovernmental organizations and NGOs). It is also useful as a text for water and sanitation practices in developing countries.
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admin 21 Mar, 2013
TweetAbstract Access to safe drinking water is important as a health and development issue at national, regional, and local levels. About one billion people do not have healthy drinking water. More than six million people (about two million children) die because of diarrhea which is caused by polluted water. Developing countries pay a high cost [...]
Access to safe drinking water is important as a health and development issue at national, regional, and local levels. About one billion people do not have healthy drinking water. More than six million people (about two million children) die because of diarrhea which is caused by polluted water. Developing countries pay a high cost to import chemicals including polyaluminium chloride and alum. This is the reason why these countries need low-cost methods requiring low maintenance and skill. The use of synthetic coagulants is not regarded as suitable due to health and economic considerations. The present study was aimed to investigate the effects of alum as coagulant in conjunction with bean, sago, and chitin as coagulants on the removal of color, turbidity, hardness, and Escherichia coli from water. A conventional jar test apparatus was employed for the tests. The study was taken up in three stages, initially with synthetic waters, followed by testing of the efficiency of coagulants individually on surface waters and, lastly, testing of blended coagulants. The experiment was conducted at three different pH conditions of 6, 7, and 8. The dosages chosen were 0.5, 1, 1.5, and 2 mg/l. The results showed that turbidity decrease provided also a primary E. coli reduction. Hardness removal efficiency was observed to be 93% at pH 7 with 1-mg/l concentration by alum, whereas chitin was stable at all the pH ranges showing the highest removal at 1 and 1.5mg/l with pH 7. In conclusion, using natural coagulants results in considerable savings in chemicals and sludge handling cost may be achieved.
Alum; Chitin; Sago; Bean; Coagulation; Turbidity
The explosive growth of the world’s human population and subsequent water and energy demands have led to an expansion of standing surface water . Nowadays, the concern about contamination of aquatic environments has increased, especially when water is used for human consumption. About one billion people do not have healthy drinking water. More than six million people (about two million children) die because of diarrhea which is caused by polluted water[2,3].
In most of the cases, surface water turbidity is caused by the clay particles, and the color is due to the decayed natural organic matter. Generally, the particles that determine the turbidity are not separated by settling or through traditional filtration. Colloidal suspension stability in surface water is also due to the electric charge of particle surface. Thus, there is great importance in either the development of more sophisticated treatments or the improvement of the current ones .
The production of potable water from most raw water sources usually entails the use of a coagulation flocculation stage to remove turbidity in the form of suspended and colloidal material. This process plays a major role in surface water treatment by reducing turbidity, bacteria, algae, color, organic compounds, and clay particles. The presence of suspended particles would clog filters or impair disinfection process, thereby dramatically minimizing the risk of waterborne diseases [5,6].
Many coagulants are widely used in conventional water treatment processes, based on their chemical characteristics. These coagulants are classified into inorganic, synthetic organic polymers, and natural coagulants . Alum has been the most widely used coagulant because of its proven performance, cost effectiveness, relatively easy handling, and availability. Recently, much attention has been drawn on the extensive use of alum. Aluminum is regarded as an important poisoning factor in dialysis encephalopathy. Aluminum is one of the factors which might contribute to Alzheimer’s disease [7-9]. Alum reaction with water alkalinity reduces water pH and its efficiency in cold water [10,11]. However, some synthetic organic polymers such as acrylamide have neurotoxicity and strong carcinogenic effect [8,12].
In addition, the use of alum salts is inappropriate in some developing countries because of the high costs of imported chemicals and low availability of chemical coagulants . This is the reason why these countries need low-cost methods requiring low maintenance and skill.
For these reasons, and also due to other advantages of natural coagulants/flocculants over chemicals, some countries such as Japan, China, India, and the United States have adopted the use of natural polymers in the treatment of surface water for the production of drinking water . A number of studies have pointed out that the introduction of natural coagulants as a substitute for metal salts may ease the problems associated with chemical coagulants.
Natural macromolecular coagulants are promising and have attracted the attention of many researchers because of their abundant source, low price, multi-purposeness, and biodegradation[11,14,15]. Okra, rice, and chitosan are natural compounds which have been used in turbidity removal [16-18]. The extract of the seeds has been mentioned for drastically reducing the amount of sludge and bacteria in sewage .
In view of the above discussion, the present work has been taken up to evaluate the efficiency of various natural coagulants on the physico-chemical contaminant removal of water. To date, most of the research has been concentrated on the coagulant efficiencies in synthetic water, but in this study, we move ahead making an attempt to test the efficiency of the natural coagulants on surface water. The efficiencies of the coagulants as stated by  might alter depending on many factors: nature of organic matter, structure, dimension, functional groups, chemical species, and others.
Natural coagulants and their preparation
Sago is a product prepared from the milk of tapioca root. Its botanical name is ‘Manihot esculentaCrantz syn. M. utilissima’. Hyacinth bean with botanical name Dolichos lablab is chosen as another coagulant. Both the coagulants were used in the form of powders (starches). Starch consists mainly of a homopolymer of α-D-glucopyranosyl units that comes in two molecular forms, linear and branched. The former is referred to as amylose and the latter as amylopectin . These have the general structure as per  (Figure 1) .
Figure 1. General structure of amylose and amylopectin.
The third coagulant was chitin ([C8H13O5N]n), which is a non-toxic, biodegradable polymer of high molecular weight. Like cellulose, chitin is a fiber, and in addition, it presents exceptional chemical and biological qualities that can be used in many industrial and medical applications. The two plant originated coagulants were taken in the form of powder or starch. Chitin was commercially procured.
The first stage included testing the efficiency of the four coagulants on the synthetic waters. Synthetic waters with turbidity of 70 and 100 nephelometric turbidity units (NTU) were prepared with fuller’s earth in the laboratory and were used in this part of the study. The experiment was carried out using a jar test apparatus. The experiments were conducted in duplicates to eliminate any kind of error. Efficiency was evaluated by determination of reduction in turbidity of both the synthetic samples.
In the second stage of the experiment, the individual coagulants were evaluated for their efficiency on the surface waters. The water samples for this stage and the preceding stage were collected from the surface reservoir, Mudasarlova, located at a distance of 5 km from the Environmental Monitoring Laboratory, GITAM University, where the experiments were carried out. This is the reservoir which serves as a source of domestic water for the nearby residents.
Care was taken while collecting the samples so that a representative sample is obtained. All samples were collected in sterile plastic containers. The samples were transported to the laboratory, and all the experiments were conducted within a duration of 24 h. The physical parameters like temperature and color were noted at the point of sample collection. The water samples were analyzed for the following parameters pre- and post-treatment with the coagulants (Table 1).
Table 1. Physico-chemical parameters tested (stage II)
The coagulants were tested at various concentrations like 0.5, 1, 1.5, and 2 mg/l at three pH ranges of 6, 7, and 8.
The results obtained from the second stage of the study have encouraged us to further extend the study in terms of blended coagulants. The blending of coagulants was taken up from the fact that alum was the most widely used coagulant, and hence, it was taken as one part. The remaining combinations were 2, 3, 4, and 5 parts of the natural coagulants, i.e., 1:2, 1:3, 1:4, and 1:5.
Testing of the following parameters was adopted for evaluating the efficiency of the blended coagulants (pre- and post-coagulation) (Table 2). All the analysis has been performed as per the standard methods given by APHA, 2005 .
Table 2. Physico-chemical parameters tested (stage III)
E. coli presence
The E. coli bacterial presence and absence were determined in the pre- and post-coagulated water using H2S strip bottle. The water sample was filled into the bottle and allowed to stand for 24 h at room temperature. After 24 h, the water sample was observed for color change; black color change indicates the presence of E. coli.
Coagulant actions onto colloidal particles take place through charge neutralization of negatively charged particles. If charge neutralization is the predominant mechanism, a stochiometric relation can be established between the particles’ concentration and coagulant optimal dose.
In the initial stage of the experiment, the coagulants were tested against synthetic turbid samples with 70 and 100 NTU. According to Figure 2a,b, the optimum dosage of alum was observed to be 1mg/l for both the turbid samples, and the optimum pH is observed to be 7.
Figure 2. Turbidity removal efficiency of alum with initial turbidities of (a) 100 and (b) 70 NTU.
It is understood from Figure 3a,b that the optimum dosage for chitin as coagulant is 1.5 mg/l (turbidity to 40 NTU) for 100 NTU, whereas not much difference was observed between pH 7 and 8 for both the turbid samples. The optimum pH is observed to be 7 for both 70 and 100 NTU samples.
Figure 3. Turbidity removal efficiency of chitin with initial turbidities of (a) 100 and (b) 70 NTU.
Figure 4a,b exemplifies the trends of sago on the turbidity removal of the synthetic solutions. It is observed that sago was effective at both 1 and 1.5 mg/l (turbidity reduced to 50 and 45 NTU, respectively) for 100 NTU solution, and the efficiency was stable at pH 7 and 8.
Figure 4. Turbidity removal efficiency of sago with initial turbidities of (a) 100 and (b) 70 NTU.
Figure 5a,b illustrates the effect of bean on the synthetic turbid samples and turbidity removal. It is observed that bean was effective at 1mg/l (turbidity reduced to 55 NTU) for 100 NTU solution, and the efficiency was stable at pH 7 and 8.
Figure 5. Turbidity removal efficiency of bean with initial turbidities of (a) 100 and (b) 70 NTU.
Implications from the stage 1 experiment articulate that the coagulants are quite stable at the pH ranges tested; hence, in the proceeding experiments, all the three pH ranges were considered. In the second stage of experiment, the environmental samples from the surface water source were collected and tested for the removal of turbidity and other chemical parameters. The dosages were the same as the previous stage. The results are graphically represented as shown in Figures 6, 7,8, 9.
Figure 6. Turbidity removal efficiency of individual coagulants.
Figure 7. Total hardness removal efficiency of individual coagulants.
Figure 8. Calcium hardness removal efficiency of individual coagulants.
Figure 9. Chloride removal efficiency of coagulants.
The turbidity removal efficiencies of the individual coagulants are depicted in Figure 6 wherein there was a broad variation among the pH ranges. The maximum reduction was observed with 1 mg/l (87%) of bean at pH 6 followed by 1 mg/l (82%) sago at the same pH. At pH 7, the maximum efficiency was shown by bean with 1.5 mg/l dosage (85.37%) followed by bean and sago with 1 (82.49%) and 1.5 mg/l (82.49%), respectively. Removal efficiencies of 41.46% and 36.59% were reported by 1 mg/l of bean and sago, respectively, at pH 8. The minimum reductions are not reported as there was a negative competence of the coagulants at different doses and pH variations. It can be observed from the graph that there was an increase in the turbidity of the water at these dosages like with 2 g of chitin the turbidity removal was −19.51. In the entire study, the best results were obtained with total hardness removal wherein no negative competence was reported as shown in Figure 7. The utmost removal was observed with 0.5-mg/l (97.67%) sago at pH 7. At pH 6, it was (90.70%) with 1.5 mg/l of bean. At pH 8, the reduction was (93.02%) with 0.5 mg/l of alum. Apart from these, the general observation was that all the coagulants were effective in an average removal of 65% total hardness at all pH variations and doses. The tracking for the least efficiency has showed chitin at pH 6 with 2-mg/l dose (34.88%).
The calcium hardness removal efficiencies are directly proportional with the total hardness removal; the highest removal was recorded by chitin (93.33%) at pH 7 with 1.5-mg/l dose as shown in Figure 8. Removal of 90% is at pH 8 and 7 with 0.5-mg/l alum and 1-mg/l chitin, respectively. Minimum effectiveness was observed by chitin (6.67%) at pH 6 with 2-mg/l dose. On an average, the removal competence was more than 60% with all coagulants at doses at all the pH conditions.
Figure 8 illustrates the chloride removal efficiency of the coagulants tested. The average competence was observed to be 40%. The maximum competence was noted at pH 7 by chitin (83.64%) at 1.5 mg/l followed by sago (81.82%) at 1 mg/l. Indeed at pH 7, the removal was observed to be superior as a whole. Similarly, pH has shown inferior effectiveness in the amputation of chloride. The remarkable point that was noted is that at pH 8, where the removal was superior, the increase in doses of sago and bean (1.5 and 2 mg/l) has shown a depressing outcome.
With the results obtained from the second stage experimentation, the study was carried forward for the evaluation of blended coagulants. From the literature, it was understood that blended coagulants show improved competence than that of the individual ones.
The regular test of turbidity was substituted with conductivity to establish a relation and test the difference with these parameters. The conductivity diminution was observed to be preeminent at the ratio of 1:2 of all the blended coagulants 26.12%, 26.00%, and 21.35% with alum/bean, alum/chitin, and alum/sago, respectively. The highest reduction was observed with alum/sago at pH 8 with 1:2 ratio (32.28%) (Figure 10).
Figure 10. Conductivity removal efficiency of blended coagulants.
The total hardness reduction trend of the blended coagulants was recorded as follows: at pH 7, all combinations of alum/bean have resulted in negative competence. Amputation of 100% was observed with alum/chitin and alum/sago at 1:2 and 1:4 and 1:5 doses, respectively (Figure 11). The overall competence of the alum/chitin and alum/sago were registered to be more than 80%. The calcium hardness efficiencies of the blended coagulants were similar to that of the total hardness. The highest removal efficiency was shown by alum/chitin with 1:5 ratio at pH 7 (Figure 12).
Figure 11. Total hardness removal efficiency of blended coagulants.
Figure 12. Calcium hardness removal efficiency of blended coagulants.
As said earlier, the turbidity was replaced by color determination taking into account the fact that turbidity is directly related to the color. pH 7 has been remarkably effective in the highest removal of color from the water. The blended coagulant alum/sago was found to be very effective with 98% to 100% reduction in color at all the ratios of dosage (Figure 13). The blended coagulants alum/chitin and alum/sago were relatively successful at an average rate of 80% decline in the color at almost all ratios of dosage at pH 7 and 8.
Figure 13. Color removal efficiency of blended coagulants.
Alum/sago blend has a noteworthy effect on the removal of chloride from the water samples in which no negative result was noted. The highest reduction was observed with alum/chitin with dose of 1:5 (85.71%) at pH 7. Indeed, pH 7 can be optimized as perfect pH for this blend as all the ratios of dosages were quite efficient in the removal of chloride (Figure 14).
Figure 14. Chloride removal efficiency of blended coagulants.
Although many studies have used synthetic water in the experiments, this work chose to use raw water collected directly from the surface source. Therefore, it is important to consider that the natural compounds may cause variations in their composition, which interfere in the treatment process. All those factors are taken into account when evaluating the obtained results.
The characteristics of the superficial water used in this study are observed as that the water used has apparent color, turbidity, solids, and amount of compounds with a relatively high absorption in UV (254 nm). It is noticeable that the water has high turbidity and color.
The effectiveness of alum, commonly used as a coagulant, is severely affected by low or high pH. In optimum conditions, the white flocs were large and rigid and settled well in less than 10 min. This finding is in agreement with other studies at optimum pH [24,25]. The optimum pH was 7 and was similar to the obtained results by Divakaran . At high turbidity, a significant improvement in residual water turbidity was observed. The supernatant was clear after about 20-min settling. Flocs were larger and settling time was lower. The results showed that above optimum dosage, the suspensions showed a tendency to restabilize.
The effectiveness of the chitin in the present study in the removal of various contaminants with varied pH individually and also in blended form can be traced to the explanation from the literature that chitin has been studied as biosorbent to a lesser extent than chitosan; however, the natural greater resistance of the former compared to the last, due to its greater crystallinity, could mean a great advantage. Besides, the possibility to control the degree of acetylation of chitin permits to enhance its adsorption potential by increasing its primary amine group density. Recent studies regarding the production of chitin-based biocomposites and its application as fluoride biosorbents have demonstrated the potential of these materials to be used in continuous adsorption processes. Moreover, these biocomposites could remove many different contaminants, including cations, organic compounds, and anions .
Chitosan has high affinity with the residual oil and excellent properties such as biodegradability, hydrophilicity, biocompability, adsorption property, flocculating ability, polyelectrolisity, antibacterial property, and its capacity of regeneration in many applications . It has been used as non-toxic floccules in the treatment of organically polluted wastewater .
The effects of coagulation process on hardness are observed for varying levels of hardness, which resulted in significant decrease of hardness removal. The study correlates with the results obtained by , wherein they had a maximum hardness removal of 84.3% by chitosan in low turbid water with initial hardness of about 204 mg/l as CaCO3.
Several experiments were carried out to determine the comparative performance of chitosan on E. coli in different turbidities. E. coli negative is present in the chitin-treated waters in all of the turbidities. The conclusive evidence was found for the negative influence of chitosan on E. coli. The regrowth of E. coli was not observed in the experiments after 24 h, which was similar to the observations by .
As far as sago is considered, the starch was effective both individually and as blended coagulant. Unlike polyaluminium chloride, the efficiency of the natural coagulants is not affected by pH. The pH increased their efficiency, which is one of the advantages of natural coagulants. The principle behind the efficiency of the sago from the literature can be stated as follows: Sago starch is a natural polymer that is categorized as polyelectrolyte and can act as coagulant aid. Coagulant aid can be classified according to the ionization traits, which are the anions, cations, and amphoteric (with dual charges). Bratskaya et al.  mentioned that among the three groups, cation polymer is normally used to remove adsorbed negatively charged particles by attracting the adsorbed particles through electrostatic force. They discovered that anion polymer and those non-ionized cannot be used to coagulate negatively charged particles.
The chemical oxygen demand (COD) reduction is influenced by the concentration of sago used; the lower the concentration the better the removal of the COD. Using less than 1.50 g L-1, better COD reduction is observed. At this low concentration, settling time did not influence the COD reduction. Similarly, concentration of sago used at lower than 1.50 g L-1 reduced the turbidity in less than 15 min of settling time. Sago concentration higher than 1.50 g L-1 increased the turbidity; however, settling time has an influence on the turbidity reduction at higher sago concentrations. This pattern is congruent with the COD removal .
The sago starch-graft-polyacrylamide (SS-g-PAm) coagulants were found to achieve water turbidity removal up to 96.6%. The results of this study suggest that SS-g-PAm copolymer is a potential coagulant for reducing turbidity during water treatment .
At its optimum concentration, D. lablab seed powder does not affect the pH of the water. Total and calcium hardness remained almost constant and were within acceptable levels according to World Health Organization standards for drinking water. Moreover, coagulation of medium to high turbidity water with D. lablab seed powder with the finest grain size reduced turbidity further. The best performance of the finest seed powder could be due to its large total surface area, whereby most of the water-soluble proteins are at the solid–liquid interface during the extraction process as stated by Gassenschmidtet al. . This might have increased the concentration of active coagulation polymer in the extract, which improved the coagulation process. The coagulant extract from seeds has shown antimicrobial activity in the comparative culture test, which was also observed in the study of Tandonet al. .
D. lablab demonstrated the best performance with turbid water, in which a turbidity removal efficiency of 87% was observed. The restabilization of destabilized colloidal particles, which was associated with higher residual turbidities, occurred at dosages above the optimum. It is commonly observed that particles are destabilized by small amounts of hydrolyzing metal salts and that optimum destabilization corresponds with neutralization of the particles’ charge. Larger amounts of coagulants cause charge reversal so that the particles become positively charged and, thus, restabilization occurs, which results in elevated turbidity levels . It has also been observed that the reduction in turbidity is associated with significant improvements in bacteriological quality. The effect of natural coagulants on turbidity removal and the antimicrobial properties against microorganisms may render them applicable for simultaneous coagulation and disinfection of water for rural and peri-urban people in developing countries .
It is observed that blended coagulants gave utmost efficiency as compared to the traditional alum coagulants. Here in this blending process, we reduce the alum dose up to 80%; thus, we reduce the drawbacks of the alum. Also, we can reduce the cost of the treatment using the natural coagulants instead of the traditional coagulant.
E. coli is the best coliform indicator of fecal contamination from human and animal wastes. E. colipresence is more representative of fecal pollution because it is present in higher numbers in fecal material and generally not elsewhere in the environment . Results showed the absence of E. coli increases with increasing time. A greater percentage of E. coli was eliminated in higher turbidities. The aggregation and, thus, removal of E. coli was directly proportional to the concentration of particles in the suspension. Chitosan and other natural coagulants showed antibacterial effects of 2 to 4 log reductions.
Antimicrobial effects of water-insoluble chitin and coagulants were attributed to both its flocculation and bactericidal activities. A bridging mechanism has been reported for bacterial coagulation by chitosan . Especially with reference to chitosan, molecules can stack on the microbial cell surface, thereby forming an impervious layer around the cell that blocks the channels, which are crucial for living cells . On the other hand, cell reduction in microorganisms, such as E. coli, occurred without noticeable cell aggregation by chitosan.
This indicates that flocculation was not the only mechanism by which microbial reduction occurred. It was found that when samples were stored during 24 h, regrowth of E. coli was not observed for all turbidities. It should be noted that the test water contained no nutrient to support regrowth of E. coli, and chitosan is not a nutrient source for it. Another experiment was designed to check the effect of alum alone. Regrowth of E. coli was not observed for unaided alum after 24 h. The number of E. coli after resuspension of sediment reached to the initial numbers after 24 h and showed that it cannot be inactivated by alum. Such findings have been previously reported by Bina.
Access to clean and safe drinking water is difficult in rural areas of India. Water is generally available during the rainy season, but it is muddy and full of sediments. Because of a lack of purifying agents, communities drink water that is no doubt contaminated by sediment and human feces. Thus, the use of natural coagulants that are locally available in combination with solar radiation, which is abundant and inexhaustible, provides a solution to the need for clean and safe drinking water in the rural communities of India. Use of this technology can reduce poverty, decrease excess morbidity and mortality from waterborne diseases, and improve overall quality of life in rural areas.
The application of coagulation treatment using natural coagulants on surface water was examined in this study. The surface water was characterized by a high concentration of suspended particles with a high turbidity. At a varied range of pH, the suspended particles easily dissolved and settled along with the coagulants added. Research has been undertaken to evaluate the performance of natural starches of sago flour, bean powder, and chitin to act as coagulants individually and in blended form. In all three cases, the main variable was the dosage of the coagulant. The study shows that natural characteristics of starch and other coagulants can be an efficient coagulant for surface water but would need further study in modifying it to be efficient to the maximum. Thus, it can be concluded that the blended coagulants are the best which give maximum removal efficiency in minimum time.
It is chitin and chitosan which can readily be derivatized by utilizing the reactivity of the primary amino group and the primary and secondary hydroxyl groups to find applications in diversified areas. In this work, an attempt has been made to increase the understanding of the importance and effects of chitin at various doses and pH conditions, upon the chemical and biological properties of water. In view of this, this study will attract the attention of academicians and environmentalists.
This is an Open Access article distributed under the terms of the Creative Commons Attribution License ( http://creativecommons.org/licenses/by/2.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.
International Journal of Energy and Environmental Engineering 2012, 3:29 doi:10.1186/2251-6832-3-29
Department of Environmental Studies, GITAM Institute of Science, GITAM University, Visakhapatnam, Andhra Pradesh 530045, India
The electronic version of this article is the complete one and can be found online at:http://www.journal-ijeee.com/content/3/1/29
||24 May 2012
||30 July 2012
||5 October 2012
© 2012 Vara; licensee BioMed Central Ltd.
admin 26 Feb, 2013
TweetDrinking water quality standards describes the quality parameters set for drinking water. Despite the truism that every human on this planet needs drinking water to survive and that water can contain many harmful compounds, there are no universally recognized and accepted international standards for drinking water. Even where standards exist and are applied, the permitted [...]
Drinking water quality standards describes the quality parameters set for drinking water. Despite the truism that every human on this planet needs drinking water to survive and that water can contain many harmful compounds, there are no universally recognized and accepted international standards for drinking water. Even where standards exist and are applied, the permitted concentration of individual constituents may vary by as much as ten times from one set of standards to another.
Many developed countries specify standards to be applied in their own country. In Europe, this includes the European Drinking Water Directive and in the USA the United States Environmental Protection Agency (EPA) establishes standards as required by the Safe Drinking Water Act. For countries without a legislative or administrative framework for such standards, the World Health Organization publishes guidelines on the standards that should be achieved. China adopted its own drinking water standard GB3838-2002 (Type II) enacted by Ministry of Environmental Protection in 2002.
Where drinking water quality standards do exist, most are expressed as guidelines or targets rather than requirements, and very few water standards have any legal basis or are subject to enforcement. Two exceptions are the European Drinking Water Directive and the Safe Drinking Water Act in the USA, which require legal compliance with specific standards.
In Europe, this includes a requirement for member states to enact appropriate local legislation to mandate the directive in each country. Routine inspection and, where required, enforcement is enacted by means of penalties imposed by the European Commission on non-compliant nations.
Countries with guideline values as their standards include Canada which has guideline values for a relatively small suite of parameters, New Zealand where there is a legislative basis but water providers have to make “best efforts” to comply with the standards in Australia.
Range of standards
Although drinking water standards are frequently referred to as if they are simple lists of parametric values, standards documents also specify the sampling location, sampling methods, sampling frequency, analytical methods and laboratory accreditation AQC. In addition, a number of standards documents also require calculation to determine whether a level exceeds the standard, such as taking an average. Some standards give complex, detailed requirements for the statistical treatment of results, temporal and seasonal variations, summation of related parameters, and mathematical treatment of apparently aberrant results.
A parametric value in this context is most commonly the concentration of a substance, e.g. 30 mg/l of Iron. It may also be a count such as 500 E. coli per litre or a statistical value such as the average concentration of copper is 2 mg/l. Many countries not only specify parametric values that may have health impacts but also specify parametric values for a range of constituents that by themselves are unlikely to have any impact on health. These include colour, turbidity, pH and the organoleptic (aesthetic) parameters (taste and odor).
It is possible and technically acceptable to refer to the same parameter in different ways that may appear to suggest a variation in the standard required. For example, nitrite may be measured as nitrite ion or expressed as N. A standard of “Nitrite as N” set at 1.4 mg/l equals a nitrite ion concentration of 4.6 mg/l – an apparent difference of nearly threefold.
Drinking water quality standards in Australia have been developed by the Australian Government National Health and Medical Research Council (NHMRC) in the form of the Australian Drinking Water Guidelines. These guidelines provide contaminant limits (pathogen, aesthetic, organic, inorganic and radiological) as well as guidance on applying limits for the management of drinking water in Australian drinking water treatment and distribution.
European Union standards
The following parametric standards are included in the Drinking Water directive and are expected to be enforced by appropriate legislation in every country in the European Union. Simple parametric values are reproduced here but in many cases the original directive also provides caveats and notes about many of the values given.
• Acrylamide 0.10 μg/l
• Antimony 5.0 μg/l
• Arsenic 10 μg/l
• Benzene 1.0 μg/l
• Benzo(a)pyrene 0.010 μg/l
• Boron 1.0 mg/l
• Bromate 10 μg/l
• Cadmium 5.0 μg/l
• Chromium 50 μg/l
• Copper 2.0 mg/l
• Cyanide 50 μg/l
• 1,2-dichloroethane 3.0 μg/l
• Epichlorohydrin 0.10 μg/l
• Fluoride 1.5 mg/l
• Lead 10 μg/l
• Mercury 1.0 μg/l
• Nickel 20 μg/l
• Nitrate 50 mg/l
• Nitrite 0.50 mg/l
• Pesticides 0.10 μg/l
• Pesticides – Total 0.50 μg/l
• Polycyclic aromatic hydrocarbons 0.10 μg/l Sum of concentrations of specified compounds;
• Selenium 10 μg/l
• Tetrachloroethene and Trichloroethene 10 μg/l Sum of concentrations of specified parameters
• Trihalomethanes — Total 100 μg/l Sum of concentrations of specified compounds
• Vinyl chloride 0.50 μg/l
United States standards
In the USA, the federal legislation controlling drinking water quality is the Safe Drinking Water Act (SDWA) which is implemented by the EPA, mainly through state or territorial primacy agencies. States and territories must implement rules at least as stringent as EPA’s to retain primary enforcement authority (primacy) over drinking water. Many states also apply their own state-specific standards which may be more rigorous or include additional parameters. Standards set by the EPA in the USA are not international standards since they apply to a single country. However, many countries look to the USA for appropriate scientific and public health guidance and may reference or adopt USA standards.
World Health Organisation (WHO) guidelines
The WHO guidelines include the following recommended limits on naturally occurring constituents that may have direct adverse health impact:
• Arsenic 0.010 mg/l
• Barium 10μg/l
• Boron 2400μg/l
• Chromium 50μg/l
• Fluoride 1500μg/l
• Selenium 40μg/l
• Uranium 30μg/l
For man-made pollutants potentially occurring in drinking water, the following standards have been proposed:
• Cadmium 3μg/l
• Mercury 6μg/l For inorganic mercury
• Benzene 10μg/l
• Carbon tetrachloride 4μg/l
• 1,2-Dichlorobenzene 1000μg/l
• 1,4-Dichlorobenzene 300μg/l
• 1,2-Dichloroethane 30μg/l
• 1,2-Dichloroethene 50μg/l
• Dichloromethane 20μg/l
• Di(2-ethylhexyl)phthalate 8 μg/l
• 1,4-Dioxane 50μg/l
• Edetic acid 600μg/l
• Ethylbenzene 300 μg/l
• Hexachlorobutadiene 0.6 μg/l
• Nitrilotriacetic acid 200μg/l
• Pentachlorophenol 9μg/l
• Styrene 20μg/l
• Tetrachloroethene 40μg/l
• Toluene 700μg/l
• Trichloroethene 20μg/l
• Xylenes 500μg/l
Comparison of parameters
The following table provides a comparison of a selection of parameters concentrations listed by WHO, the European Union, EPA and Ministry of Environmental Protection of China.
” indicates that no standard has been identified by editors of this article and ns indicates that no standard exists. μg/l -> Micro grams per litre or 0.001 ppm, mg/L -> 1 ppm or 1000 μg/l (Text made available under the Creative Commons Attribution-ShareAlike License: original found here: http://en.wikipedia.org/wiki/Drinking_water_quality_standards
World Health Organization
||50 μg/l (Cr6)
||10 mg/L (as N)
||10 mg/L (as N)
||1 mg/L (as N)
||0.10 μg/ l
|Pesticides — Total
|Polycyclic aromatic hydrocarbons l
|Tetrachloroethene and Trichloroethene
admin 21 Feb, 2013
TweetContent Table Recent Papers in Adsorption and Ion Exchange Processes Magnetic ion exchange resin treatment for drinking water production Removal of radiocobalt from EDTA-complexes using oxidation and selective ion exchange Ammonium removal from anaerobic digester effluent by ion exchange A hybrid ion exchange-nanofiltration (HIX-NF) process for energy efficient desalination of brackish/seawater Adsorption kinetics and isotherm [...]
Magnetic ion exchange resin treatment for drinking water production
Journal of Water Supply: Research and Technology—AQUA Vol 58 No 1 pp 41–50 © IWA Publishing 2009 doi:10.2166/aqua.2009.081
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B. Sani, E. Basile, L. Rossi and C. Lubello
Department of Civil and Environmental Engineering, University of Florence, Via S. Marta 3, I-50139, Florence, Italy Tel.: +39 55 479 6317 E-mail: email@example.com
Publiacqua SpA, Via Villamagna 39, I-50126, Florence, Italy
Italian drinking water treatment plants (DWTP) generally use chlorine-based chemicals to achieve the oxidation/disinfection phases of their treatment trains. The main problem related to the application of such disinfectants consists in the formation of disinfection by-products (DBPs) as a result of the reaction with organic substances in the water. Italian regulations set very strict limits for the maximum concentration of chlorine DBPs and, for many DWTPs, the compliance with such a regulation is difficult. Non-oxidative pre-treatments, able to remove organic substances from the water prior to chlorination, could be a suitable solution to overcome this problem. These treatments could increase the water quality, decrease the oxidant demand and, hence, reduce the formation of DBPs. This paper presents an experimental investigation of ion exchange processes for the dissolved organic carbon (DOC) removal by using MIEX® resin. The process was studied as a pre-treatment on raw river water. The DOC removal efficiency and the effects on downstream processes of the treatment train were evaluated.
Removal of radiocobalt from EDTA-complexes using oxidation and selective ion exchange
Water Science & Technology—WST Vol 60 No 4 pp 1097–1101 © IWA Publishing 2009 doi:10.2166/wst.2009.458
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L. K. Malinen, R. Koivula and R. Harjula
Laboratory of Radiochemistry, Department of Chemistry, University of Helsinki, P.O. Box 55 (A. I. Virtasen aukio 1), FI-00014, Finland E-mail: firstname.lastname@example.org; email@example.com; firstname.lastname@example.org
Methods for the removal of radiocobalt from an ethylenediaminetetraacetic acid (EDTA) complex of Co(II) (aqueous solution containing 10 mM Co(II) and 10 mM or 50 mM EDTA traced with 57Co) are presented. The studies examined a combination of different oxidation methods and the sorption of 57Co on a selective inorganic ion exchange material, CoTreat. The oxidation methods used were ultraviolet (UV) irradiation with and without hydrogen peroxide (H2O2), as well as ozonation alone or in combination with UV irradiation. Also, the possible contribution of Degussa P25 TiO2 photocatalyst to degradation of EDTA was studied. The best results for the equimolar solution of Co(II) and EDTA were achieved by combining ozonation, UV irradiation, Degussa P25 TiO2 and CoTreat, with approximately 94% sorption of 57Co. High values for the 57Co sorption were also achieved by ozonation (~88%) and UV irradiation (~90%) in the presence of CoTreat and Degussa P25 TiO2. A surplus of EDTA over Co(II) was also tested using 10 mM Co(II) and 50mM EDTA. Only a slight decrease, to ~88% sorption of 57Co, was detected compared to the value (~90%) obtained with 10 mM EDTA.
Ammonium removal from anaerobic digester effluent by ion exchange
Water Science & Technology—WST Vol 60 No 1 pp 201–210 © IWA Publishing 2009 doi:10.2166/wst.2009.317
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T. Wirthensohn, F. Waeger, L. Jelinek and W. Fuchs
Department of IFA-Tulln, Institute for Environmental Biotechnology, University of Natural Resources and Applied Life Sciences—Vienna, Konrad Lorenz Strasse 20, 3430 Tulln, Austria E-mail: email@example.com; firstname.lastname@example.org; email@example.com
Department of Power Engineering, Faculty of Environmental Technology, Institute of Chemical Technology, Technicka 5, 166 28 Prague 6, Czech Republic E-mail: Ludek.Jelinek@vscht.cz
The effluent of a 500 kW biogas plant is treated with a solid separation, a micro filtration and a reverse osmosis to achieve nutrient recovery and an effluent quality which should meet disposal quality into public water bodies. After the reverse osmosis, the ammonium concentration is still high (NH4-N = 467 mg/l), amongst other cations (K+=85 mg/l; Na+=67 mg/l; Mg2 + =0.74 mg/l; Ca2 + =1.79 mg/l). The aim of this study was to remove this ammonium by ion exchange. Acidic gel cation exchange resins and clinoptilolite were tested in column experiments to evaluate their capacity, flow rates and pH. Amberjet 1,500 H was the most efficient resin, 57 BV of the substrate could be treated, 1.97 mol NH4-N/l resin were removed. The ammonium removal was more than 99% and the quality of the effluent was very satisfactory (NH4-N < 2 mg/l). The breakthrough of the observed parameters happened suddenly, the order was sodium—pH—ammonium—potassium. The sharp increase of the pH facilitates the online control, while the change in conductivity is less significant. A regeneration with 3 bed volumes of 2 M HCl recovered 91.7% of the original cation exchange capacity.
A hybrid ion exchange-nanofiltration (HIX-NF) process for energy efficient desalination of brackish/seawater
Water Science & Technology: Water Supply—WSTWS Vol 9 No 4 pp 369–377 © IWA Publishing 2009 doi:10.2166/ws.2009.634
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S. Sarkar and A. K. SenGupta
Department of Civil and Environmental Engineering, Lehigh University, Fritz Engineering Laboratory, 13 E Packer Avenue, Bethlehem PA, 18015, USA E-mail: firstname.lastname@example.org;email@example.com
This study reports a new hybrid ion exchange-nanofiltration (HIX-NF) process for desalination of sea and brackish water that can attain significant energy economy over the conventional membrane-based pressure driven processes. In this hybrid process, an ion exchange step converts monovalent chloride ions of saline water to divalent sulfate ions and the resulting solution, having a reduced osmotic pressure than the feed, is desalinated using a nanofiltration (NF) membrane. The sulfate rich reject stream from the NF process is used to regenerate the anion exchanger. Results validate that NF membranes can desalinate sodium sulfate solution at a much lower transmembrane pressure compared to RO membranes as well as yield a higher permeate flux. The sulfate-chloride selectivity of the anion exchangers plays important role in sustainability of the process. Laboratory studies have revealed that a single type of anion exchanger cannot sustain the process for saline water with different salt concentrations. However, anion exchangers with different sizes of amine functional groups (e.g. quaternary-, tertiary-, secondary- and primary amine) hold the promise that the process can be tailored to achieve sustainability. Laboratory studies have validated the basic premise of the hybrid process including greater than two times less energy requirement than RO process for the same feed water and same permeate recovery condition.
Adsorption kinetics and isotherm characteristics of selected endocrine disrupting compounds on activated carbon in natural waters
Water Science & Technology: Water Supply—WSTWS Vol 9 No 1 pp 51–58 © IWA Publishing 2009 doi:10.2166/ws.2009.063
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A. Assoumani, L. Favier-Teodorescu and D. Wolbert
Ecole Nationale Supérieure de Chimie de Rennes,CNRS, UMR 6226, Avenue du Général Leclerc, CS 50837, 35700, Rennes Cedex 4, France E-mail: firstname.lastname@example.org
Bisphenol A (BPA) and ethynylestradiol (EE2), two representative endocrine disrupting compounds (EDCs), were tested for their adsorbabilities onto two powdered activated carbons (PACs). The main aim of the study was to create a prediction tool for the determination of the EDCs adsorbabilities at low ng.L-1 level. Single solute solution adsorption isotherms at high concentrations, for prediction purposes, and low concentrations, for verification of the prediction, were performed for one EDC/PAC couple. Over the whole range of concentration, results showed that the Langmuir-Freundlich model better suits the adsorption phenomenon than the Freundlich or Langmuir model. Kinetics experiments were carried out on the same EDC/PAC couple. HSDM modelling of single solute adsorption kinetics at high concentration allowed determining the kinetic coefficients kf and Ds; both were shown to dominate the mass transfer mechanism. Competitive adsorption isotherms at high and low concentrations showed that downward extrapolation of low concentration adsorption capacities from solely high concentration information results in acceptable error compared to the total range isotherm. The IAST-EBC approach combined with the Langmuir-Freundlich single solute model, for the target compound, and the Langmuir model, for the EBC, appears as an acceptable global model.
Influence of hybrid coagulation-ultrafiltration pretreatment on trace organics adsorption in drinking water treatment
Journal of Water Supply: Research and Technology—AQUA Vol 58 No 3 pp 170–180 © IWA Publishing 2009 doi:10.2166/aqua.2009.071
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S. Müller and W. Uhl
Institute of Urban Water Management (ISI), Chair of Water Supply Engineering, Technische Universität Dresden, Dresden, 01062, Germany Tel.: +49-(0)351-46333126 Fax: +49-(0)351-46337204 E-mail: email@example.com
The treatment of raw water by hybrid coagulation-ultrafiltration was investigated. Coagulation-ultrafiltration removed high molecular weight organics, preferentially humics. Adsorption of the trace compound cis-1,2-dichloroethene, present in raw water, on granular activated carbon was improved considerably as compounds competing for adsorption space had been removed. This was shown in isotherms and breakthrough curves. Aeration during filtration did not affect membrane performance as expressed in permeability. However, aeration in the submerged membrane container resulted in a release of organic matter from the flocs, which resulted in higher concentrations of dissolved organic carbon in the filtrate.
Phosphorus adsorption on water treatment residual solids
Journal of Water Supply: Research and Technology—AQUA Vol 58 No 1 pp 1–10 © IWA Publishing 2009 doi:10.2166/aqua.2009.017
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Meaghan K. Gibbons, Md. Maruf Mortula and Graham A. Gagnon
Department of Civil and Resource Engineering, Dalhousie University, Halifax, Nova Scotia, B3J 1X1, Canada Tel.: +1 902 494 3268 Fax:+1 902 494 3108 E-mail: firstname.lastname@example.org
Department of Civil Engineering, American University of Sharjah, Sharjah, PO Box, 26666, UAE
The treatment and disposal of water treatment plant residual solids has become an increasingly important environmental priority for drinking water utilities. This study examines water treatment residual solids (WTRSs) from four North American water treatment plants to determine the role that coagulant types play in phosphate adsorption by the residual solids. In total, two alum residual solids (one solid from a plant that has a raw water with low alkalinity and one solid from a plant that has a raw water with high alkalinity), one lime residual solid and one ferric residual solid were used in batch adsorption experiments with deionized water at a pH of 6.2±0.2 and secondary municipal wastewater effluent at a pH of 6.8. Langmuir isotherm modeling showed that ferric residuals had the highest adsorptive capacity for phosphate (Qmax=2,960 mg/kg), followed by lime (Qmax=1,390 mg/kg) and alum (Qmax=1,110 mg/kg and 1,030 mg/kg) for adsorption experiments with P-spiked deionized water. Of the two alum residuals, the residual with a higher weight percent of metal oxides had a higher adsorptive capacity. The ferric residuals were less affected by competing species in the wastewater effluent, while the lime and alum residuals had a higher rate of phosphate removal from the deionized water compared to the wastewater effluent. Overall, ferric water treatment residuals were the best adsorbent for phosphate adsorption, followed by lime and alum residuals.
Influence of surface chemistry and structure of activated carbon on adsorption of fulvic acids from water solution
Water Science & Technology—WST Vol 60 No 2 pp 441–447 © IWA Publishing 2009 doi:10.2166/wst.2009.344
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L. A. Savchyna, I. P. Kozyatnyk, T. V. Poliakova and N. A. Klymenko
Institute of Colloid Chemistry and Chemistry of Water, Ukrainian National Academy of Sciences, 42 Vernadsky Avenue, Kiev 03680, Ukraine E-mail: email@example.com
The adsorption of fulvic acids (FA) from aqueous solutions on activated carbon (AC) with different characteristics of surface chemical state has been investigated. To characterize the adsorbability of FA with complex fractional composition, a method of estimation of modified Freundlich equation constants was employed, and “conventional component” was used to evaluate the change in Gibbs free adsorption energy. It has been shown that change in activated carbon surface energy in-homogeneity due to oxidation leads mainly to a decrease in the adsorption energy of fulvic acids and to an increase of the concentration range of the conventional portion of the low adsorbable fraction. Decrease in the adsorption energy of organic substrate may result in higher degree of spontaneous bioregeneration of activated carbon and hence in its longer life in the processes of FA solutions filtration.
Synthesis of carboxylated chitosan and its adsorption properties for cadmium (II), lead (II) and copper (II) from aqueous solutions
Water Science & Technology—WST Vol 60 No 2 pp 467–474 © IWA Publishing 2009 doi:10.2166/wst.2009.369
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K. L. Lv, Y. L. Du and C. M. Wang
Department of Chemistry, Lanzhou University, Lanzhou 730000, China E-mail: firstname.lastname@example.org
Carboxylated chitosan (CKCTS) was prepared for the removal of Cd(II), Pb(II), and Cu(II) from aqueous solutions. The effects of experimental parameters such as pH value, initial concentration, contact time and temperature on the adsorption were studied. From the results we can see that the adsorption capacities of Cd(II), Pb(II), and Cu(II) increase with increasing pH of the solution. The kinetic rates were best fitted to the pseudo-second-order model. The adsorption equilibrium data were fitted well with the Langmuir isotherm, which revealed that the maximum adsorption capacities for monolayer saturation of Cd(II), Pb(II), and Cu(II) were 0.555, 0.733 and 0.827 mmol/g, respectively. The adsorption was an exothermic process.
Water Science & Technology—WST Vol 59 No 2 pp 303–310 © IWA Publishing 2009 doi:10.2166/wst.2009.865
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M. A. Hossain, H. Furumai and F. Nakajima
Institute of Water and Flood Management, Bangladesh University of Engineering and Technology, Dhaka, 1000, Bangladesh E-mail: email@example.com; firstname.lastname@example.org
Research Center for Water Environment Technology, The University of Tokyo, 7-3-1 Hongo, Bunkyo-ku, Tokyo, 113-8656, Japan E-mail: email@example.com
Environmental Science Center, The University of Tokyo, 7-3-1 Hongo, Bunkyo-ku, Tokyo, 113-0033, Japan E-mail- firstname.lastname@example.org
Accumulation of heavy metals at elevated concentration and potential of considerable amount of the accumulated heavy metals to reach the soil system was observed from earlier studies in soakaways sediments within an infiltration facility in Tokyo, Japan. In order to understand the competitive adsorption behaviour of heavy metals Zn, Ni and Cu in soil, competitive batch adsorption experiments were carried out using single metal and binary metal combinations on soil samples representative of underlying soil and surface soil at the site. Speciation analysis of the adsorbed metals was carried out through BCR sequential extraction method. Among the metals, Cu was not affected by competition while Zn and Ni were affected by competition of coexisting metals. The parameters of fitted ‘Freundlich’ and ‘Langmuir’ isotherms indicated more intense competition in underlying soil compared to surface soil for adsorption of Zn and Ni. The speciation of adsorbed metals revealed less selectivity of Zn and Ni to soil organic matter, while dominance of organic bound fraction was observed for Cu, especially in organic rich surface soil. Compared to underlying soil, the surface soil is expected to provide greater adsorption to heavy metals as well as provide greater stability to adsorbed metals, especially for Cu.
admin 11 Feb, 2013
TweetStudy of Physico-Chemical Characteristics of Wastewater in an Urban Agglomeration in Romania Abstract This study investigates the level of wastewater pollution by analyzing its chemical characteristics at five wastewater collectors. Samples are collected before they discharge into the Danube during a monitoring campaign of two weeks. Organic and inorganic compounds, heavy metals, and biogenic compounds [...]
Study of Physico-Chemical Characteristics of Wastewater in an Urban Agglomeration in Romania
This study investigates the level of wastewater pollution by analyzing its chemical characteristics at five wastewater collectors. Samples are collected before they discharge into the Danube during a monitoring campaign of two weeks. Organic and inorganic compounds, heavy metals, and biogenic compounds have been analyzed using potentiometric and spectrophotometric methods. Experimental results show that the quality of wastewater varies from site to site and it greatly depends on the origin of the wastewater. Correlation analysis was used in order to identify possible relationships between concentrations of various analyzed parameters, which could be used in selecting the appropriate method for wastewater treatment to be implemented at wastewater plants.
Sources of wastewater in the selected area are microindustries (like laundries, hotels, hospitals, etc.), macroindustries (industrial wastewater) and household activities (domestic wastewater). Wastewater is collected through sewage systems (underground sewage pipes) to one or more centralized Sewage Treatment Plants (STPs), where, ideally, the sewage water is treated. However, in cities and towns with old sewage systems treatment stations sometimes simply do not exist or, if they exist, they might not be properly equipped for an efficient treatment. Even when all establishments are connected to the sewage system, the designed capacities are often exceeded, resulting in a less efficient sewage system and occasional leaks.
Studies of water quality in various effluents revealed that anthropogenic activities have an important negative impact on water quality in the downstream sections of the major rivers. This is a result of cumulative effects from upstream development but also from inadequate wastewater treatment facilities. Water quality decay, characterized by important modifications of chemical oxygen demand (COD), total suspended solids (TSSs), total nitrogen (TN), total phosphorous (TP), iron (Fe), nickel (Ni), copper (Cu), zinc (Zn), lead (Pb), and so forth  are the result of wastewater discharge in rivers. Water-related environmental quality has been shown to be far from adequate due to unknown characteristics of wastewater . Thus an important element in preventing and controlling river pollution by an effective management of STP is the existence of reliable and accurate information about the concentrations of pollutants in wastewater. Studies of wastewater in Danube basins can be found, for instance, in central and eastern European countries, but we are not aware of extensive studies of wastewater quality at regional/national level in Romania.
This paper analyses the chemical composition of wastewater at several collectors/stations in an important Romanian city, Galati, before being discharged into natural receptors, which in this case are the Danube and Siret Rivers. No sewage treatment existed when the monitoring campaign took place, except the mechanical separation. The study presented here is part of a larger project aiming at establishing the best treatment technology of wastewater at each station. Presently this project is in the implementation stage at all stations. Possible relationships between concentrations of various chemical residues in wastewater and with pollution sources are also investigated. The study is based on daily measurements of chemical parameters at five city collectors in Galati, Romania, during a two-week campaign in February 2010.
2.1. Location of Sampling Sites
Galati-Braila area is the second urban agglomeration in Romania after Bucharest, which is located in Romania at the confluence of three major rivers: Danube, Siret, and Prut. The wastewater average flow is about 100000 m3/day . The drainage system covers an area of 2300 ha, serving approximately 99% of the population (approximately 300000 habitants). The basic drainage system is very old, dating back to the end of the 19th century, and was extended along with the expansion of the city due to demographic and industrial evolution. There are several collectors that collect wastewater and rainwater from various areas with very different characteristics, according to the existing water-pipe drainage system. There is no treatment at any station, except for simple mechanical separation. However, industrial wastewater is pretreated before being discharged in the city system. The five wastewater collectors are denoted in the following as S 1 , S 2 , … , S 5. Four of them discharge in the Danube River and the fifth discharges in the Siret River (which is an affluent of Danube River). Figure 1 shows the distribution of the monitoring sites and highlights the type of collecting area (domestic, industrial, or mixed). For the sake of brevity, these stations will be named in the present paper as “domestic,” “mixed,” and “industrial” stations, according to the type of collected wastewater. The mixture between domestic and industrial water at the two mixed collectors is the result of changes in city planning and various transformations of small/medium enterprises.
Figure 1: Monitoring sampling sites of wastewater from Galati city.
Technical details about each collector/station can be found in Table 1. The first station, S1, collects 10% of the total quantity of wastewater. A high percentage of the water collected at this station comes from domestic sources from the south part of the city (more than 96%). Station S2 collects 64% of the total daily flow of wastewater, out of which 30% comes from domestic sources and the rest (70%) is industrial. Most of the industrial sources in this area are food-production units (milk, braid, wine) while the domestic sources include 20 schools, 4 hospitals, and important social objectives. Station S3 is located in the old part of the city and collects 5% of the total wastewater and has domestic sources. At the fourth station, S4, 11% of the quantity of wastewater is collected from domestic (70%) and industrial (30%) sources. The last collector, S5, collects wastewater from the industrial area of the city, where the most important objectives are a shipyard, metallurgical, and mechanical plants and transport stations.
Table 1: Characteristics of collectors S 1 , … , S 5.
2.2. Physico-Chemical Parameters and Methods of Analysis
The physico-chemical parameters which were measured are the following:(i)pH;(ii)chemical oxygen demand (COD) and dissolved oxygen (DO);(iii)nutrients such as nitrate (N-NO3) and phosphate (P-PO4) (these were included due to their impact on the eutrophication phenomenon);(iv)metals such as aluminum (Al+3), soluble iron (Fe+2), and cadmium (Cd+2).
The pH and DO were determined in situ using a portable multiparameter analyzer. Other chemical parameters such as COD, metals and nutrients were determined according to the standard analytical methods for the examination of water and wastewater .
The COD values reflect the organic and inorganic compounds oxidized by dichromate with the following exceptions: some heterocyclic compounds (e.g., pyridine), quaternary nitrogen compounds, and readily volatile hydrocarbons. The concentration of metals (Al+3, Cd+2, Fe+2) was determined as a result of their toxicity.
The value of pH was analyzed according to the Romanian Standard using a portable multiparameter analyzer, Consort C932.
COD parameter was measured using COD Vials (COD 25–1500 mg/L, Merck, Germany). The digestion process of 3 mL aliquots was carried out in the COD Vials for 2 h at 148°C. The absorbance level of the digested samples was then measured with a spectrophotometer at λ = 605 nm (Spectroquant NOVA 60, Merck, Germany), the method being analogous to EPA methods , US Standard Methods, and Romanian Standard Methods.
The DO parameter was analyzed according to Romanian Standard using a portable multiparameter analyzer, Consort C932.
Aluminum ions (Al+3) were determined using Al Vials (Aluminum Test 0.020–1.20 mg/L, Merck, Germany) in a way analogous to US Standard Methods. The absorbance levels of the samples were then measured with a spectrophotometer (Spectroquant NOVA 60; Merck, Germany) at λ = 550 nm. The method was based on reaction between aluminum ions and Chromazurol S, in weakly acidic-acetate buffered solution, to form a blue-violet compound that is determined spectrophotometrically. The pH of the sample must be within range 3–10. Where necessary, the pH will be adjusted with sodium hydroxide solution or sulphuric acid.
Iron concentration (Fe+2) was determined using Iron Vials (Iron Test 0.005–5.00 mg/L, Merck, Germany) and their absorbance levels were then measured using a spectrophotometer (Spectroquant NOVA 60; Merck, Germany) at λ = 565 nm. The method was based on reducing all iron ions (Fe+3) to iron ions (Fe+2). In a thioglycolate-buffered medium, these react with a triazine derivative to form a red-violet complex which is spectrophotometrically determined. The pH must be within range 3–11. Where necessary the pH was adjusted with sodium hydroxide solution or sulphuric acid.
Cadmium ions (Cd+2) were determined using Cadmium Vials (Cadmium Test 0.005–5.00 mg/L, Merck, Germany), their absorbance levels being measured with a spectrophotometer (Spectroquant NOVA 60; Merck, Germany) at λ = 525 nm. The method was based on the reaction of cadmium ions with a cadion derivative (cadion-trivial name for 1-(4-nitrophenyl)-3-(4-phenylazophenyl)triazene), in alkaline solution, to form a red complex that is determined spectrophotometrically. The pH must be within the range 3–11, and, if not, the pH will be adjusted with sodium hydroxide solution or sulphuric acid.
Nitrogen content was determined using Nitrate Vials (Nitrate Cell test in seawater 0.10–3.00 mg/L NO3-N or 0.4–13.3 mg/L N O3 −, Merck, Germany). The method being based on the reaction of nitrate ions with resorcinol, in the presence of chloride, in a strongly sulphuric acid solution, to form a red-violet indophenols dye that is determined spectrophotometrically. The absorbance levels of the samples were then measured with a spectrophotometer (Spectroquant NOVA 60; Merck, Germany) at λ = 500 nm.
Phosphorous content was determined using Phosphate Vials (Phosphate Cell Test 0.5–25.0 mg/L PO4-P or 1.5–76.7 mg/L P O4 − 3, Merck, Germany) with a method that was analogous to the US Standard Methods . The method was based on the reaction of orthophosphate anions, in a sulphuric solution, with ammonium vanadate and ammonium heptamolybdate to form orange-yellow molybdo-vanado-phosphoric acid that is determined spectrophotometrically (“VM” method). The absorbance levels of the samples were then measured with a spectrophotometer (Spectroquant NOVA 60; Merck, Germany) at λ = 410 nm.
All results were compared with standardized levels for wastewater quality found in accordance with European Commission Directive  and Romanian law .
3. Results and Discussion
3.1. The Acidity (pH)
The results for pH for all the investigated five collectors are shown in Figure 2.
Figure 2: Daily variation of pH at all sites.
Generally, the wastewater collected at the monitored sites is slightly alkaline. The pH varies between 6.8 and 8.3—average value 7.82—thus the pH values are within the accepted range for Danube River according to the Romanian law, which is between 6.5 and 9.0. The pH variation is relatively similar at collectors S1–S4 (domestic and/or mixed domestic-industrial contribution). Lower pH values are observed at S5, which is dominated by industrial wastewater, originating from major enterprises and heavy industry. However, these values are not too low, since usually pH values for industrial wastewater are smaller than 6.5.
A significant decrease in the pH value was observed during the 8th day of the analyzed period at each station. Interestingly, a heavy snowfall took place at that particular time, thus the decrease could be attributed to the mixing between wastewater and a high quantity of low pH water, resulted from the melting of snow . One could speculate that the snowfall, which has an acidic character, might have affected the pH of the wastewater through “run off” phenomena.
No other snowfall took place during the monitoring campaign, thus no definite conclusion can be drawn for a possible relationship between pH and snowfalls.
3.2. Results for Chemical Oxygen Demand (COD)
Detection of COD values in each sampling site of wastewater is presented in Figure 3.
Figure 3: Daily variation of COD at all sites.
All COD values are higher than the maximum accepted values (125 mg O2/L) of the Romanian Law . Both organic and inorganic compounds have an effect on urban wastewater’s oxidability since COD represents not only oxidation of organic compounds, but also the oxidation of reductive inorganic compounds. That means some inorganic compounds interfere with COD determination through the consumption of C r2O7 − 2. Two different behaviors can be observed, which are associated with the type of the collected wastewater as follows.(i)The first group consists of stations S2, S4 and S5 where the wastewater has an important industrial component. At these stations, COD values are approximately between 150 and 300 mg O2/L, smaller, for instance, than COD values found by in the raw wastewater produced by an industrial coffee plant where COD values were between 4000 and 4600 mg O2/L. Also, the temporal variation of COD values at all three stations is similar with no significant deviations from the average value, which is about 250 mg O2/L. Interestingly, the lowest COD level can be seen, on the average, at S5, which has the highest percentage of industrial wastewater. The second group comprises the “domestic” stations S1 and S3. The COD levels are higher, with values of 500 mg O2/L or more. Also, the variability is clearly higher than at the industrial-type stations. No clear association between the variations at the two sites can be seen. A peak in COD was measured in the 14th day of the study at site S1 (1160 mg O2/L). Since S1 is a domestic type station, it is unlikely that some major discharge led to such a high variation of COD. Unfortunately, no other information exists that might indicate a possible cause for this increase.
3.3. Results for Dissolved Oxygen (DO)
The amount of DO, which represents the concentration of chemical or biological compounds that can be oxidized and that might have pollution potential, can affect a sum of processes that include re-aeration, transport, photosynthesis, respiration, nitrification, and decay of organic matter. Low DO concentrations can lead to impaired fish development and maturation, increased fish mortality, and underwater habitat degradation . No standards are given by Romanian or European Law for DO in wastewater. The DO values for the analyzed wastewater at all five sites are shown in Figure 4.
Figure 4: Daily variation of DO at all sites.
Concentration of DO varies at all sampling sites and has values between 0.96 (at S2) and 11.33 (at S4) mg O2/L with a mean value of 6.39 mg O2/L. These are clearly higher than DO values measured, for instance, in surface natural waters in China, where the Taihu watershed had the lowest DO level (2.70 mg/L), while in other rivers DO varied from 3.14 to 3.36 mg O2/L . On the other hand, such high values of DO (9.0 mg O2/L) could be found, for instance, in the Santa Cruz River , who argued that discharging industry and domestic wastewater induced serious organic pollution in rivers, since the decrease of DO was mainly caused by the decomposition of organic compounds. Extremely low DO content (DO < 2 mg O2/L) usually indicates the degradation of an aquatic system .
The DO levels vary similarly for all selected sampling sites. The DO levels cover a wide range, with a minimum value of 1.0 mg O2/L at S1 and S3 and a maximum value of 11.33 mg O2/L at S4. There is a drop in DO at all stations, observed is in the 8th day of the monitoring interval, which coincides with the day when a similar decrease in pH took place. The lowest values of DO are observed for S1, one of the two “domestic” stations. It is interesting to note that DO at S5 is low although the wastewater here comes only from industry sources.
The variation of Al+3, Fe+2, and Cd+2 concentrations in wastewater are shown in Figures 5, 6, and 7. Al+3 concentrations (Figure 5) were mostly within the 0.05–0.20 mg/L range at all the sampling sites. However, during the beginning and the end of the monitoring campaign, Al+3 concentration at station S2 is high (reaching even 0.65 mg/L), nonetheless below the limit imposed by the Romanian law, which is 5 mg/L . The fact that in the beginning of the time interval, the concentration of Al+3 is high at two neighboring stations (S1 and S2) suggests that some localized discharge affecting both runaway and waste water, might have happened in the southern part of the city, which led to the increase of Al+3concentration in the collected wastewater. This is supported by the fact that the concentration gradually decreases at S2.
Figure 5: Daily variation of Al at all sites.
Figure 6: Daily variation of Fe at all sites.
Figure 7: Daily variation of Cd at all sites.
The variation of Fe+2 concentrations is shown in Figure 6. Fe+2 concentration is within the 0.07–0.4 mg/L interval, below 5.0 mg/L, which is the maximum accepted value of the Romanian law . Two higher values were observed at S2 and S4 (both with industrial component) during the third and fourth days of the monitoring campaign.
Besides Al+3 and Fe+2, concentrations of Cd+2 were determined and the variations at the five stations are shown in Figure 7. Cd+2 is a rare pollutant, originating from heavy industry. Leakages in the sewage systems can also lead to Cd+2. Except for two days, Cd+2 varies between 0.005 and 0.04 mg/L. The two high values of 0.11 mg/L were observed in the first and fourth days at S5, which collects industrial wastewater. However, Cd+2 concentrations do not exceed the maximum accepted values of the Romanian law  for the monitoring interval which is 0.2 mg/L.
Water systems are very vulnerable to nitrate pollution sources like septic systems, animal waste, commercial fertilizers, and decaying organic matter . Important quantities of nutrients, which are impossible to be removed naturally, can be found in rivers and this leads to the eutrophication of natural water (like Danube River). As a result, an increase in the lifetime of pathogenic microorganisms is expected. Measurement of nutrient (different forms of nitrogen (N) or phosphorous (P)) variations in domestic wastewater is strongly needed in order to maintain the water quality of receptors . Nitrogen by nitrate (Figure 8) and phosphorous by phosphate (Figure 9) are considered as representative for nutrients.
Figure 8: Daily variation of N-NO3 at all sites.
Figure 9: Daily variation of P-PO4 at all sites.
Figure 8 shows that N-NO3 concentrations vary, on the average, between 0 and 5.0 mg/L.
At all four stations with a domestic component, S1, S2, S3 and S4, the concentration of N-NO3 is low (between 0 and 1.5 mg/L) and the daily variation is relatively similar at all sites. Noticeable drops of the N-NO3 concentration are observed at all stations in the 8th day of the monitoring interval, coinciding with pH (Figure 2) and DO strong variations (Figure 4). This supports the conclusion that the heavy snowfall recorded at that period had an important impact on wastewater quality most likely due to the runoff joining the sewage system.
The behavior of N-NO3 clearly differs at station S5, which collects only industrial wastewater. Significantly higher values of N-NO3, ranging from 2.0 to 5.0 mg/L, were detected. However, the mean concentration of N-NO3 remained below the maximum concentration given by the Romanian law . Obviously, if treatment stations have to be set up, the priority for this particular nutrient component should concentrate on stations where industrial wastewater is collected.
Another nutrient that was analyzed for our study was orthophosphate expressed by phosphorous. The P-PO4 concentration varies, on the average, between 1.0 and 6.0 mg/L (Figure 9). For this component, concentrations are higher at domestic stations, S1 and S3, than at the other three stations. P-PO4 is expected to increase in domestic wastewater because of food, more precisely meat, processing, washing, and so forth. The lowest values were observed at S5, which has a negligible domestic component. Peaks in the P-PO4 concentration are observed at S1. Interestingly enough, P-PO4 temporal variations correlated pretty well at stations S2, S4, and S5 (which collect industrial wastewater). Unlike most of the other analyzed compounds, for which the concentrations were within the accepted ranges, the maximum level of P-PO4 is exceeded at all five collectors. Both Romanian law and the European law stipulate 2.0 mg/L total phosphorous for 10000–100000 habitants, and for more than 100000 habitants (as in Galati City’s case) 1.0 mg/L total phosphorus. Interestingly, domestic stations seem to require more attention with respect to the quality of water then industrial stations.
Our results regarding the variation and levels of the analyzed parameters are grouped below as the following.(1)The values of pH are within the accepted range for Danube, and their daily variations are relatively similar for both domestic and mixed wastewater. Significantly smaller pH values were measured in the wastewater with a high industrial load. A clear minimum was observed at all sites in the 8th day of the monitoring period, when a heavy snowfall took place. One could speculate that the snowfall, which has an acidic character, might have affected the pH of the wastewater through “run off” phenomena. However, a clear connection cannot be established relying on one event only.(2)The COD level clearly depends on the type of wastewater. Higher values were observed for domestic wastewater, while “pure” industrial wastewater has the lowest COD. This might be explained by the fact that industrial wastewater benefits from some treatment before being discharged into the city sewage system. However, COD does exceed the maximum accepted values according to the Romanian law  at all sites thus additional treatment is required at all stations.(3)Concentrations of all analysed metals, Al+3, Cd+2 and Fe+2, are within the limit of the Romanian law. No association with the type of wastewater could be inferred. Isolated peaks could not be linked with any specific polluting factors, except for Cd+2, for which accidental concentration increases are observed for pure industrial wastewater.(4)The level of P-PO4, one of the two nutrients that were analyzed, was high at all stations; however, the highest concentrations are associated with domestic loads.(5)Opposingly, the N-NO3 level is the highest, by far, in wastewater with a high industrial contribution.
3.6. Possible Relationships between Various Parameters
The experimental results have shown that some parameters might be related and that their behavior greatly depends on the type of collected wastewater. Differences between the behavior of physico-chemical parameters at the domestic sites (S1 and S3), on one hand, and at the other sites, on the other, was observed. Pearson correlation coefficients have been calculated between all parameters at all the selected five sites and corresponding significances. Although most of correlations were not significant, some interesting connections between various parameters at sites with similar characteristics were revealed. Table 2 shows correlation coefficients between various parameters for all five stations. Significant correlations at different types of stations are denoted as follows: italicized fonts for domestic stations, boldface italicized fonts for the industrial station and boldface fonts for mixed stations.
Table 2: Correlation coefficients calculated for station S1 to S5. Significant correlations at each type of stations are identified as follows: boldface italicized fonts for industrial station (S5), italicized fonts for domestic stations (S1 and S3) and boldface fonts for mixed stations (S2 and S4).
An important relationship seems to exist between pH and N-NO3 at all stations except for the industrial wastewater collecting site, S5 (i.e., at all stations collecting wastewater resulting from domestic activities). Similarly, pH correlates well with DO at all stations except the industrial one.
COD correlates with two metals, Cd+2 and soluble Fe+2, which is expected , but only at S1 and S3, where the daily variations of the concentration for these two metals (Cd+2 and soluble Fe+2) were similar.
No conclusion can be drawn for the industrial wastewater collector that was analyzed, where both positive and negative correlations were observed. The lack of correlation between the two metals and COD at the industrial wastewater collectors suggests that other processes, that alter the chemical equilibrium between the two chemical compounds, must be taken into account. For example some metals are complexed by organic compounds that are present in the water and the pH values can influence these phenomena.
DO correlates with pH and N-NO3 at all four sampling stations with domestic component (S1–S4) but the relationship vanish at S5 (industrial). There is also a negative correlation between DO and Fe+2 and Cd+2 only for domestic wastewater, which is expected because of the natural oxidation of metals. The correlation vanishes at the other three stations which collect wastewater from industrial areas.
Heavy metals, Fe+2 and Cd+2 correlate only at domestic stations and no relationships can be defined to link the concentration of Al+3 with other components.
The P-PO4 variation is linked to the variation of soluble Fe+2 at the two stations that collect domestic wastewater. Interestingly, these two elements exist together in reductive domestic systems because these are dominated by proteins, lipids, degradation products. This relationship disappears at the other stations, where the industrial load is significant. The other metals, Al+3, seems to be linked with P-PO4at stations S5 and S2, which collect wastewater with the highest industrial load. No link is observed for the rest of stations and for Cd+2 which can be explained by a higher probability of iron (II) orthophosphate to form in wastewater compared to Al+3 or Cd+2 orthophosphates.
Positive correlations can also be seen between P-PO4 and COD for all sampling sites except S1, where the relationship is still positive but less significant. The other nutrient, N-NO3, is anticorrelated with COD but only at S3 and is well correlated with pH and DO at all four stations with domestic component. The only exception is station S5, which collects mostly industrial wastewater.
Concluding, positive correlations were observed between the following parameters.(1)pH and N-NO3 everywhere except “purely” industrial water.(2)COD and soluble Fe+2 at domestic stations.(3)DO and pH, on the one hand, and DO and N-NO3 at domestic stations.(4)P-PO4 and soluble Fe+2 at domestic stations.(5)P-PO4 and COD everywhere, which, taking into account the high level of P-PO4 at domestic stations, might suggest that one important contributor to water quality degradation are household discharges.(6)Al+3 and P-PO4.
In the present paper we have analyzed the daily variation of several physico-chemical parameters of the wastewater (pH, COD, DO, Al+3, Fe+2, Cd+2, N-NO3, and P-PO4) at five collectors that have been characterized as domestic, industrial and mixed, according to the type of collecting area. Different results have been obtained for domestic and industrial wastewater. Most of the chemical parameters are within accepted ranges. Nevertheless, their values as well as their behavior depend significantly on the type of collected wastewater.
The overall conclusion is that wastewater with a high domestic load has the highest negative impact on water quality in a river. On the other hand, industrial wastewater brings an important nutrient load, with potentially negative effect on the basins where it is discharged. Our results suggested that meteorological factors (snow) might modify some characteristics of wastewater, but a clear connection cannot be established relying on one event only.
Significantly smaller pH values were measured in the wastewater with a high industrial load. The COD level clearly depends on the type of wastewater. Higher values were observed for wastewater with domestic sources, while “pure” industrial wastewater has the lowest COD. This might be explained by the fact that industrial wastewater benefits from some treatment before being discharged into the city sewage system. COD does exceed the maximum accepted values according to the Romanian law at all sites thus additional treatment is required at all stations. Accidental increases of Cd+2 concentrations are observed for pure industrial wastewater. The highest concentrations of P-PO4 are associated with domestic loads. Opposing, the N-NO3 level is clearly the highest in wastewater with a high industrial contribution.
Correlation analysis has been used in order to identify possible relationships between various parameters for wastewater of similar origin.
Positive correlations between various physico-chemical parameters exist for the domestic wastewater (DO, pH and N-NO3, on the one hand, and P-PO4, COD and soluble Fe+2, on the other hand). Except for two cases, these relationships break when the industrial load is high. Some of the existing correlations are expected as discussed above, thus any removal treatment should be differentiated according to the type of collector, before discharging it into the natural receptors in order to be costly efficient. Correlations between DO and COD and nutrient load suggest that the most important threat for natural basins in the studied area, are domestic sources for the wastewater.
The different percentages of industrial and domestic collected wastewater vary at each station, which has a clear impact on concentrations of the selected chemical components. Our results show that domestic wastewater has a higher negative impact on water quality than wastewater with a high industrial load, which, surprisingly, seems to be cleaner. This might be related to the fact that most industries are forced, by law, to apply a pretreatment before discharging wastewater into the city sewage system. Industrial wastewater affects the nutrient content of natural water basins. Although the time period was relatively short, our study identified specific requirements of chemical treatment at each station. An efficient treatment plan should take into account the type of wastewater to be processed at each station. Results presented here are linked with another research topic assessing the level of water quality in the lower basin of the Danube before and after implementing the complete biochemical treatment plants.
The work of Catalin Trif was supported by Project SOP HRD-EFICIENT 61445/2009.
Copyright © 2012 Paula Popa et al. This is an open access article distributed under the Creative Commons Attribution License, which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited – original found here: http://www.hindawi.com/journals/tswj/2012/549028/
admin 30 Jan, 2013
TweetUltraviolet germicidal irradiation (UVGI) is a disinfection method that uses ultraviolet (UV) light at sufficiently short wavelength to kill microorganisms. It is used in a variety of applications, such as food, air and water purification. UVGI uses short-wavelength ultraviolet radiation that is harmful to microorganisms. It is effective in destroying the nucleic acids in these [...]
A low pressure mercury vapor discharge tube floods the inside of a biosafety cabinet with shortwave UV light when not in use, sterilizing microbiological contaminants from irradiated surfaces.
Ultraviolet germicidal irradiation (UVGI) is a disinfection method that uses ultraviolet (UV) light at sufficiently short wavelength to kill microorganisms. It is used in a variety of applications, such as food, air and water purification. UVGI uses short-wavelength ultraviolet radiation that is harmful to microorganisms. It is effective in destroying the nucleic acids in these organisms so that their DNA is disrupted by the UV radiation, leaving them unable to perform vital cellular functions.
The wavelength of UV that causes this effect is rare on Earth as the atmosphere blocks it. Using a UVGI device in certain environments like circulating air or water systems creates a deadly effect on micro-organisms such as pathogens, viruses and molds that are in these environments. Coupled with a filtration system, UVGI can remove harmful microorganisms from these environments.
The application of UVGI to disinfection has been an accepted practice since the mid-20th century. It has been used primarily in medical sanitation and sterile work facilities. Increasingly it was employed to sterilize drinking and wastewater, as the holding facilities were enclosed and could be circulated to ensure a higher exposure to the UV. In recent years UVGI has found renewed application in air sanitizing.
UV has been a known mutagen at the cellular level for more than one-hundred years. The 1903 Nobel Prize for Medicine was awarded to Niels Finsen for his use of UV against lupus vulgaris, tuberculosis of the skin.
Using ultraviolet (UV) light for drinking water disinfection dates back to 1916 in the U.S. Over the years, UV costs have declined as researchers develop and use new UV methods to disinfect water and wastewater. Currently, several states have developed regulations that allow systems to disinfect their drinking water supplies with UV light.
Ultraviolet light is electromagnetic radiation with wavelengths shorter than visible light. UV can be separated into various ranges, with short range UV (UVC) considered “germicidal UV.” At certain wavelengths UV is mutagenic to bacteria, viruses and other microorganisms. At a wavelength of 2,537 Angstroms (254 nm) UV will break the molecular bonds within micro-organismal DNA, producing thymine dimers in their DNA thereby destroying them, rendering them harmless or prohibiting growth and reproduction. It is a process similar to the UV effect of longer wavelengths (UVB) on humans, such as sunburn or sun glare. Microorganisms have less protection from UV and cannot survive prolonged exposure to it.
A UVGI system is designed to expose environments such as water tanks, sealed rooms and forced air systems to germicidal UV. Exposure comes from germicidal lamps that emit germicidal UV electromagnetic radiation at the correct wavelength, thus irradiating the environment. The forced flow of air or water through this environment ensures the exposure.
The effectiveness of germicidal UV in such an environment depends on a number of factors: the length of time a micro-organism is exposed to UV, power fluctuations of the UV source that impact the EM wavelength, the presence of particles that can protect the micro-organisms from UV, and a micro-organism’s ability to withstand UV during its exposure.
In many systems redundancy in exposing micro-organisms to UV is achieved by circulating the air or water repeatedly. This ensures multiple passes so that the UV is effective against the highest number of micro-organisms and will irradiate resistant micro-organisms more than once to break them down.
The effectiveness of this form of sterilization is also dependent on line-of-sight exposure of the micro-organisms to the UV light. Environments where design creates obstacles that block the UV light are not as effective. In such an environment the effectiveness is then reliant on the placement of the UVGI system so that line-of-sight is optimum for sterilization.
Sterilization is often misquoted as being achievable. While it is theoretically possible in a controlled environment, it is very difficult to prove and the term ‘disinfection’ is used by companies offering this service as to avoid legal reprimand. Specialist companies will often advertise a certain log reduction i.e. 99.9999% effective, instead of sterilization. This takes into consideration a phenomenon known as light and dark repair (photoreactivation and excision (BER) respectively) in which the DNA in the bacterium will fix itself after being damaged by UV light.
A separate problem that will affect UVGI is dust or other film coating the bulb, which can lower UV output. Therefore bulbs require annual replacement and scheduled cleaning to ensure effectiveness. The lifetime of germicidal UV bulbs varies depending on design. Also the material that the bulb is made of can absorb some of the germicidal rays.
Lamp cooling under airflow can also lower UV output, thus care should be taken to shield lamps from direct airflow via parabolic reflector. Or add additional lamps to compensate for the cooling effect.
Increases in effectiveness and UV intensity can be achieved by using reflection. Aluminium has the highest reflectivity rate versus other metals and is recommended when using UV.
Inactivation of microorganisms
The degree of inactivation by ultraviolet radiation is directly related to the UV dose applied to the water. The dosage, a product of UV light intensity and exposure time, is usually measured in microjoules per square centimeter, or alternatively as microwatt seconds per square centimeter (µW·s/cm2). Dosages for a 90% kill of most bacteria and virus range from 2,000 to 8,000 µW·s/cm2. Dosage for larger parasites such as Cryptosporidium require a lower dose for inactivation. As a result, the US EPA has accepted UV disinfection as a method for drinking water plants to obtain Cryptosporidium, Giardia or virus inactivation credits. For example, for one-decimal-logarithm reduction of Cryptosporidium, a minimum dose of 2,500 µW·s/cm2 is required based on the US EPA UV Guidance Manual published in 2006.
Weaknesses and strengths
UV water treatment devices can be used for well water and surface water disinfection. UV treatment compares favorably with other water disinfection systems in terms of cost, labor and the need for technically trained personnel for operation: deep tube wells fitted with hand pumps, while perhaps the simplest to operate, require expensive drilling rigs, are immobile sources, and often produce hard water that is found distasteful. Chlorine disinfection treats larger organisms and offers residual disinfection, but these systems are expensive because they need a special operator training and a steady supply of a potentially hazardous material. Finally, boiling water over a biomass cook stove is the most reliable treatment method but it demands labor, and imposes a high economic cost. UV treatment is rapid and, in terms of primary energy use, approximately 20,000 times more efficient than boiling.
UV disinfection is most effective for treating a high clarity purified reverse osmosis distilled water. Suspended particles are a problem because microorganisms buried within particles are shielded from the UV light and pass through the unit unaffected. However, UV systems can be coupled with a pre-filter to remove those larger organisms that would otherwise pass through the UV system unaffected. The pre-filter also clarifies the water to improve light transmittance and therefore UV dose throughout the entire water column. Another key factor of UV water treatment is the flow rate: if the flow is too high, water will pass through without enough UV exposure. If the flow is too low, heat may build up and damage the UV lamp.
In UVGI systems the lamps are shielded or are in environments that limit exposure, such as a closed water tank or closed air circulation system, often with interlocks that automatically shut off the UV lamps if the system is opened for access by human beings.
In human beings, skin exposure to germicidal wavelengths of UV light can produce sunburn and skin cancer. Exposure of the eyes to this UV radiation can produce extremely painful inflammation of the cornea and temporary or permanent vision impairment, up to and including blindness in some cases. UV can damage the retina of the eye.
Another potential danger is the UV production of ozone. Ozone can be harmful to health. The United States Environmental Protection Agency designated 0.05 parts per million (ppm) of ozone to be a safe level. Lamps designed to release UVC and higher frequencies are doped so that any UV light below 254 nm will not be released, thus ozone is not produced. A full spectrum lamp will release all UV wavelengths and will produce ozone as well as UVC, UVB, and UVA. (The ozone is produced when UVC hits oxygen (O2) molecules, and so is only produced when oxygen is present.)
UV-C radiation is able to break down chemical bonds. This leads to rapid ageing of plastics (insulations, gasket) and other materials. Note that plastics sold to be “UV-resistant” are tested only for UV-B, as UV-C doesn’t normally reach the surface of the Earth. When UV is used near plastic, rubber, or insulations care should be taken to shield said components; metal tape or aluminum foil will suffice.
A disadvantage of the technique is that water treated by chlorination is resistant to reinfection, where UVGI water must be transported and delivered in such a way as to avoid contamination.
UVGI can be used to disinfect air with prolonged exposure. Disinfection is a function of UV concentration and time, CT. For this reason, it is not as effective on moving air, when the lamp is perpendicular to the flow, as exposure times are dramatically reduced. Air purification UVGI systems can be freestanding units with shielded UV lamps that use a fan to force air past the UV light. Other systems are installed in forced air systems so that the circulation for the premises moves micro-organisms past the lamps. Key to this form of sterilization is placement of the UV lamps and a good filtration system to remove the dead micro-organisms. For example, forced air systems by design impede line-of-sight, thus creating areas of the environment that will be shaded from the UV light. However, a UV lamp placed at the coils and drainpan of cooling system will keep micro-organisms from forming in these naturally damp places.
ASHRAE covers UVGI and its applications in IAQ and building maintenance in its 2008 Handbook, HVAC Systems and Equipment in Chapter 16 titled Ultraviolet Lamp Systems. ASHRAE’s 2011 Handbook, HVAC Applications, covers ULTRAVIOLET AIR AND SURFACE TREATMENT in Chapter 60.
Ultraviolet disinfection of water consists of a purely physical, chemical-free process. UV-C radiation attacks the vital DNA of the bacteria directly. The bacteria lose their reproductive capability and are destroyed. Even parasites such as Cryptosporidia or Giardia, which are extremely resistant to chemical disinfectants, are efficiently reduced. UV can also be used to remove chlorine and chloramine species from water ; this process is called photolysis, and requires a higher dose than normal disinfection. The sterilized microorganisms are not removed from the water. UV disinfection does not remove dissolved organics, inorganic compounds or particles in the water. However, UV-oxidation processes can be used to simultaneously destroy trace chemical contaminants and provide high-level disinfection, such as the world’s largest indirect potable reuse plant in Orange County, California. That title will soon be taken by New York which is set to open the Catskill-Delaware Water Ultraviolet Disinfection Facility, by the end of 2012. A total of 56 energy-efficient UV reactors will be installed to treat 2.2 billion US gallons (8,300,000 m3) a day to serve New York City.
UV disinfection leaves no taint, chemicals or residues in the treated water. Disinfection using UV light is quick and clean.
UV tube project
The UV Tube is a design concept for providing inexpensive water disinfection to people in poor countries. The concept is based the ability of ultraviolet light to kill infectious agents by disrupting their DNA. It was initially developed under an “open source” model at the Renewable and Appropriate Energy Laboratory at the University of California, Berkeley. The form and composition of the UV Tube can vary depending on the resources available and the preferences of those building and using the device. However, certain geometric parameters must be maintained to ensure consistent performance. Several different versions of the UV Tube are currently being used in multiple locations in Mexico and Sri Lanka.
Germicidal UV is delivered by a mercury-vapor lamp that emits UV at the germicidal wavelength. Mercury vapour emits at 254 nm. Many germicidal UV bulbs use special ballasts to regulate electrical current flow to the bulbs, similar to those needed for fluorescent lights. In some cases, UVGI electrodeless lamps can be energised with microwaves, giving very long stable life and other advantages[clarification needed]. This is known as ‘Microwave UV.’
Lamps are either amalgam or medium pressure lamps. Each type has specific strengths and weaknesses.
Low-pressure UV lamps
These offer high efficiencies (approx 35% UVC) but lower power, typically 1 W/cm power density (power per unit of arc length).
Amalgam UV lamps
A high-power version of low-pressure lamps. They operate at higher temperatures and have a lifetime of up to 16,000 hours. Their efficiency is slightly lower than that of traditional low-pressure lamps (approx 33% UVC output) and power density is approx 2–3 W/cm.
These lamps have a broad and pronounced peak-line spectrum and a high radiation output but lower UVC efficiency of 10% or less. Typical power density is 30 W/cm³ or greater.
Depending on the quartz glass used for the lamp body, low-pressure and amalgam UV lamps emit light at 254 nm and 185 nm (for oxidation). 185 nm light is used to generate ozone.
The UV units for water treatment consist of a specialized low pressure mercury vapor lamp that produces ultraviolet radiation at 254 nm, or medium pressure UV lamps that produce a polychromatic output from 200 nm to visible and infrared energy. The optimal wavelengths for disinfection are close to 260 nm. Medium pressure lamps are approximately 12% efficient, whilst amalgam low pressure lamps can be up to 40% efficient. The UV lamp never contacts the water, it is either housed in a quartz glass sleeve inside the water chamber or mounted external to the water which flows through the transparent UV tube. It is mounted so that water can pass through a flow chamber, and UV rays are admitted and absorbed into the stream.
Sizing of a UV system is affected by three variables: flow rate, lamp power and UV transmittance in the water. UV manufacturers typically developed sophisticated Computational Fluid Dynamics (CFD) models validated with bioassay testing. This typically involves testing the UV reactor’s disinfection performance with either MS2 or T1 bacteriophages at various flow rates, UV transmittance and power levels in order to develop a regression model for system sizing. For example, this is a requirement for all drinking water systems in the United States per the US EPA UV Guidance Manual.:5-2
The flow profile is produced from the chamber geometry, flow rate and particular turbulence model selected. The radiation profile is developed from inputs such as water quality, lamp type (power, germicidal efficiency, spectral output, arc length) and the transmittance and dimension of the quartz sleeve. Proprietary CFD software simulates both the flow and radiation profiles. Once the 3-D model of the chamber is built, it’s populated with a grid or mesh that comprises thousands of small cubes.
Points of interest—such as at a bend, on the quartz sleeve surface, or around the wiper mechanism—use a higher resolution mesh, whilst other areas within the reactor use a coarse mesh. Once the mesh is produced, hundreds of thousands of virtual particles are “fired” through the chamber. Each particle has several variables of interest associated with it, and the particles are “harvested” after the reactor. Discrete phase modeling produces delivered dose, headless and other chamber specific parameters.
When the modeling phase is complete, selected systems are validated using a professional third party to provide oversight and to determine how closely the model is able to predict the reality of system performance. System validation uses non-pathogenic surrogates to determine the Reduction Equivalent Dose (RED) ability of the reactors. Most systems are validated to deliver 40 mJ/[cm.sup.2] within an envelope of flow and transmittance.
To validate effectiveness in drinking water systems, the methods described in the US EPA UV Guidance Manual is typically used by the U.S. Environmental Protection Agency, whilst Europe has adopted Germany’s DVGW 294 standard. For wastewater systems, the NWRI/AwwaRF Ultraviolet Disinfection Guidelines for Drinking Water and Water Reuse protocols are typically used, especially in wastewater reuse applications.
UV systems destined for drinking water applications are validated using a third party test house to demonstrate system capability, and usually a non pathogenic surrogate such as MS 2 phage or Bacillus Subtilis is used to verify actual system performance. UV manufacturers have verified the performance of a number of reactors, in each case iteratively improving the predictive models.
Ultraviolet in wastewater treatment is replacing chlorination due to the chemical’s toxic by-products. Individual wastestreams to be treated by UVGI must be tested to ensure that the method will be effective due to potential interferences such as suspended solids, dyes or other substances that may block or absorb the UV radiation.
“UV units to treat small batches (1 to several liters) or low flows (1 to several liters per minute) of water at the community level are estimated to have costs of 0.02 US$ per 1000 liters of water, including the cost of electricity and consumables and the annualized capital cost of the unit.” (WHO)
Large scale urban UV wastewater treatment is performed in cities such as Edmonton, Alberta. The use of ultraviolet light has now become standard practice in most municipal wastewater treatment processes. Effluent is now starting to be recognised as a valuable resource, not a problem that needs to be dumped. Many wastewater facilities are being renamed as water reclamation facilities, and whether the waste water is being discharged into a river, being used to irrigate crops, or injected into an aquifer for later recovery. Ultraviolet light is now being used to ensure water is free from harmful organisms.
Aquarium and pond
Ultraviolet sterilizers are often used in aquaria and ponds to help control unwanted microorganisms in the water. Continuous sterilization of the water neutralizes single-cell algae and thereby increases water clarity. UV irradiation also ensures that exposed pathogens cannot reproduce, thus decreasing the likelihood of a disease outbreak in an aquarium. UV irradiation can also have a positive impact on an Aquariums Redox balance
Aquarium and pond sterilizers are typically small, with fittings for tubing that allows the water to flow through the sterilizer on its way from a separate external filter or water pump. Within the sterilizer, water flows as close as possible to the ultraviolet light source. Water pre-filtration is critical so as to lower water turbidity which will lower UVC penetration. Many of the better UV Sterilizers have long dwell times and limit the space between the UVC source and the inside wall of the UV Sterilizer device.
UVGI is often used to disinfect equipment such as safety goggles, instruments, pipettes, and other devices. Lab personnel also disinfects glassware and plasticware this way. Microbiology laboratories use UVGI to disinfect surfaces inside biological safety cabinets (“hoods”) between uses.
Food and beverage protection
Since the FDA issued a rule in 2001 requiring that virtually all fruit and vegetable juice producers follow HACCP controls, and mandating a 5-log reduction in pathogens, UVGI has seen some use in sterilization of fresh juices such as fresh-pressed apple cider.
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