Algal control and enhanced removal in drinking waters in Cairo, Egypt

In Egypt, the majority of the population depends upon the river Nile as the main source for drinking water and for industrial and agricultural activities. However, after the completion of the Aswan High Dam in 1966, stream-flow regulation has caused significant changes to the Nile aquatic ecosystem and waterway environment (Ismail & Ramadan 1995; El-Shinnawy et al. 2000; Talling et al. 2009). This change in flow led to the retention of water masses in reservoirs behind dams and barrages in some regions within the Nile system, which provided favorable conditions and ample time for phytoplankton development. Subsequently, this altered the aquatic vegetation which is a source of nutrients for phytoplankton and algae utilization.

Since 2006, specifically in September/October, incidents of algal blooms involving the major water treatment plants all over the country have been reported by the Egyptian media. The increase in algal growth was attributed to the extended storage time of raw water on site before treatment as well as the lack of regular maintenance and pipe cleansing, an event scheduled every 6 months (Al Ahram Newspaper, 28 September 2006). Also, the prolonged residence time of water masses in the distribution system was reported to promote unexpected algal growth and microbial re-growth. Eventually, this caused an increase in algal toxins within the water distribution system during this period which was reflected by reports of at least 85 sick and one documented death in 2006 (Egyptian Minister of Health, 15 October 2006).

Generally, algae are widely dispersed throughout lakes and rivers at relatively low-population densities; however, the population is liable to rapidly grow with the increase in nutrients in surface waters (Henderson et al. 2008). This may result in dissolved oxygen depletion, high water turbidity and eventually the degradation of aquatic ecosystems (Manahan 2000; Abou El-Ella et al. 2007; Shen et al. 2011; Wang et al. 2013). Also, this may cause acute disturbance in taste, odor and induce toxins under stressful environmental conditions (Shehata et al. 2008; Heng et al. 2009; Shen et al. 2011). In a conventional water treatment plant, uncontrolled algal blooms may clog screens and rapid sand filters and hinder the drinking water supply system (Babel & Takizawa 2011; Wu et al. 2011). Alternatively, due to their microscopic size, algae may penetrate filters and reside within the domestic drinking water supply storage systems and excrete extracellular organic products that may lead to an increase in the coagulant demand, impair the final water disinfection process and promote microbial re-growth in the water distribution systems (Ma & Liu 2002; Babel & Takizawa 2011; Wu et al. 2011). Therefore, reduction of surface water algae during the initial stages of treatment is preferred to maintain acceptable final water quality (James & Conrad 2004; Henderson et al. 2008; Babel & Takizawa 2011; Wu et al. 2011). This is mainly because the removal of algal cells as suspended solids may reduce the associated soluble toxins that may impair the physico-chemical water treatment systems.

The general techniques applied for algal control in drinking water treatment systems have been discussed elsewhere (Henderson et al. 2008; Gao et al. 2010; Babel & Takizawa 2011; Shen et al. 2011; Wu et al. 2011; Devrimci et al. 2012). Of these operations, coagulation/flocculation is regarded as the main process for removal of algae as it has the ability to remove suspended solids through particle destabilization while flocculation allows for solid separation either through settling or flotation, and filtration (James & Conrad 2004; Henderson et al. 2008; Gao et al. 2010; Shen et al. 2011; Devrimci et al. 2012). However, due to their small cell size and low specific gravity, removal of algae may not be effectively achieved through using chemical coagulation alone (Devrimci et al. 2012). Therefore, chemical pre-oxidants (such as ozone, chlorine dioxide, chlorine, or permanganate) are required to enhance coagulation as they have the capacity to inactivate algal cells (Ma & Liu 2002; Chen & Yeh 2005; Henderson et al. 2008; Shehata et al. 2008; Heng et al. 2009; Gao et al. 2010). Overall, chlorine is the preferred pre-oxidant because it is a powerful water disinfectant that is safer to apply, is more economical than other algicides and is less time-consuming in application (Shen et al. 2011). Nevertheless, it was reported that algae may consume large quantities of chlorine and thus reduce the free chlorine available to control bacteria if they persist after treatment (Shen et al. 2011). On the other hand, it was reported that overdosing of pre-oxidants must be avoided as it can cause algal cell lysis and release undesirable toxins or offensive taste- and odor-associated compounds (Henderson et al. 2008; Heng et al. 2009).

In Egypt, a typical drinking water treatment system comprises stages of pre-chlorination, flocculation/coagulation with alum, filtration (14 sand and carbon filters), post-chlorination, and settling in distribution tanks (Donia 2007). In this set-up, pre-chlorination is regarded as the most important stage for the control of bacteria and algal growth (El-Dib & Rizka 1995; Donia 2007; Badawy et al. 2012); therefore, a number of studies were conducted to investigate the contribution of this stage toward the formation of disinfection by-products in the final treated waters and to set recommendations for its application (El-Dib & Rizka 1995; Donia 2007; Badawy et al. 2012). Nonetheless, in a review by Donia (2007), it was suggested that there was a need to upgrade the current treatment technologies in Egypt in order to efficiently handle the recent added loads of chemical and biological pollutants transported through surface and drinking water systems. However, there has been no investigation exploring the relationship between the raw water quality parameters and the reported algal blooms. Also, no study has been conducted to evaluate the impact of the current applied treatment procedures upon the final treated water parameters in case a combined pre-oxidation stage was applied.

The sampling site is located at El-Maadi district (Figure 1), south of the Greater Cairo Urban Region, Egypt (latitude: 29 °56′53.3″N; longitude: 31 °15′24.6″E). This location represents a major water intake point for a number of water treatment plants operating south of the capital city. Generally, the intake pipe for a water treatment plant is neither totally immersed at the bottom of the Nile nor set on the surface to avoid the intake of silt and other floating pollutants, respectively. It is common for the water treatment intake pipes for these plants to situated at an ample distance upstream from the urban district which it is to serve. However, due to the random urbanization along the river banks, the sampling location was found to be bordered to the north and east by mixed agricultural and residential areas, and to the west and south there are vast agricultural lands.

Water samples (5 L) were collected four times per month from December 2011 to November 2012. Temperature and pH of the samples were measured on site. The samples were transferred immediately to the laboratory for physical and chemical characterization according to Standard Methods(APHA 2005). The parameters of interest were measured by the following methods: turbidity: nephelometry; total dissolved solids (TDS): electrical conductivity; total alkalinity: titration (pH 4.5); [Ca²⁺] and [Mg²⁺]: Na₂EDTA titration; chlorides: argentometric method; (SO₄²⁻): turbidometry; barium chloride: spectrophotometry; PO₄³⁻: stannous chloride colorimetry; total organic carbon (TOC): persulfate-ultraviolet or heated-persulfate oxidation; silicates: silicomolybdate colorimetry. Nitrates were determined using the sodium salicylate method for nitrate determination in drinking waters (Scheiner 1974). Algal count was determined using a Sedgwick–Rafter counting chamber (APHA 2005). Residual content of Al, Fe and Mn in the final treated waters was determined by Eriochrome Cyanine R; phenanthroline, and persulfate methods, respectively (APHA 2005). An average of 4–5 determinations per parameter was undertaken, totaling 16–20 times per month.

Jar tests were conducted to evaluate algal removal by pre-oxidation followed by coagulation using a six-paddle stirrer (JLT 6, VELP Scientifica, Italy) in surface water samples at room temperature (about 20 °C). This method was used to determine the optimal oxidant as well as the coagulant demand/dose (APHA 2005).

Cluster analysis was performed using Ward's method for minimum variance and the similarity coefficient used was the square Euclidean distance function using ‘StatistiXL 1.8’ within Microsoft Excel 2007 (Microsoft^® Windows 2007). Hierarchical clustering was performed for the water chemical and physical characteristics before and after treatment and the results obtained included the distance matrix, clustering strategy report and dendrogram were utilized in data interpretation accordingly.

Table 1 shows the mean monthly chemical and physical water quality parameters obtained from December 2011 to November 2012. The data showed that a higher algal count was obtained during September to November 2012. This increase may be attributed to agricultural wastewater discharge and the accumulation of pollutants and nutrients south of the sampling point (Amer & Abd El-Gawad 2012). The TDS ranged from 221 to 314 mg/L with the highest value in winter and the lowest in June 2012, which is similar to that reported by Amer & Abd El-Gawad (2012) near the study site in 2010. Generally, it has been reported that the salinity of Nile water increased going from Aswan High Dam (160 mg/L) to Cairo (around 260 mg/L) (Wahaab & Badawy 2004; Talling 2009). This was attributed to the release of agricultural wastewaters containing excess chemicals along the river. In addition, it has been reported that higher TDS values recorded during winter for the Nile waters were due to the dam closure period, during which the amount of water released from the behind the dam was less than that in summer. On the other hand, a decrease in TDS within the Nile system was related to the increase in suspended particles during the flooding period as these particles may adsorb the dissolved solids and subsequently reduce the TDS by sedimentation (Abd El-Hady 2014). The data also revealed that the highest water turbidity was observed during the winter while the lowest was during the summer, an observation that was attributed to the winter closure period (Emara et al. 2012). High turbidity of the Nile stream was attributed to the increase in water nutrient content (Amer & Abd El-Gawad 2012), which is in agreement with the current findings.

Nile water temperatures varies from 13 °C in winter to ≈30 °C in summer (Talling 2009; Abd El-Hady 2014). This variation in temperature was reported to be an indicator of water transparency which had a direct impact upon regulating algal/phytoplankton population in these waters (Abd El-Hady 2014). In addition, higher temperatures accelerate the metabolic rate of microalgae (Shen et al. 2011). The presence of carbonates in these waters was reported to be dependent upon both the pH and salinity (TDS) (El-Ghandour et al. 1985), and the increase in alkalinity (CaCO₃) of these raw waters has been reported to promote phytoplankton growth (Abd El-Hady 2014) as a result of algal utilization of dissolved CO₂ and HCO₃ in biomass formation (Manahan 2000).

Overall, it was observed that there was an increase in Ca, Mg, SO₄, Cl, PO₄ and NO₃ content during the winter period, in accordance with the findings of Emara et al. (2012). The phosphate and nitrate contents of these waters showed a similar pattern of increase and decrease over the study period. This may be attributed to the discharge of agricultural drainage (El-Ghandour et al. 1985; Ismail & Ramadan 1995; Wahaab & Badawy 2004) as well as from surrounding industrial and urban sources (Elewa 2010). Finally, the silica content of these waters ranged between 1.58 and 4.88 mg SiO₂/L with the highest in July and the lowest in December. Shehata et al. (2008) reported that low silica levels were associated with the increase in diatoms present in these waters. This variation in silicates was also attributed to the fact that diatoms utilized dissolved silicates for the formation of skeletal materials during algal productivity (Brownlee & Taylor 2002).

Diatoms predominated over Chloropyta, Cyanophyta and Euglenophyta (Table 2). This agrees with the findings of Abd El-Hady (2014) who showed that phytoplankton in the Nile system were dominated by Bacillariophyta (diatoms) all year round. In addition, the prevalence of diatoms within the river Nile was attributed to the presence of iron and dissolved silica (Shehata et al. 2008). On the other hand, while the predominance of diatoms is regarded as an indicator of good water quality under low stress conditions, the increase in Chlorophyceae and Cyanophyceae is considered an indication of organic pollution and nutrient accumulation in surface waters (Mahapatra et al. 2010).

Generally, chlorine disinfection effectivity depends on the applied dose and contact time (Casey et al. 2012). Previous work by El-Dib & Rizka (1995) recommended, as a precaution, to monitor the applied dose and not to exceed the chlorine contact time over 30 minutes for Nile waters. This was based upon their findings that the concentration of trihalomethanes (THMs) formed increased progressively as the contact time extended from 30 to 240 minutes. In addition, low chlorine doses may readily aid in the formation of chloro-substituted by-products, while higher chlorine doses may lead to production of oxidation and cleavage products of natural organic matter (NOM) and formation of THMs.

As indicated previously, while the major concern for the use of chlorine and its derivatives in water disinfection is the formation of chlorinated by-products, the application of KMnO₄, was reported to remove soluble manganese and iron which are frequently associated with the presence of cyanobacteria in raw water (Fan et al. 2013). KMnO₄ may also serve as an alternative to pre-chlorination or other oxidants for the control of color, taste and odor, and algae and for the further control of biological growth in treatment plants. In addition, previous studies have indicated that combining chlorine with permanganate during the pre-oxidation stage has a synergistic effect upon water disinfection (Heng et al. 2008).

Aluminum sulfate, ferric chloride and ferric sulfate are the most common coagulants used in drinking water treatment (Lei et al. 2009; Matilainen et al. 2010; Sengul & Gormez 2013). The main purpose in using these coagulants is to reduce NOM which may interfere with the removal of other contaminants that cause membrane fouling, contribute to piping corrosion and act as a substrate for bacterial growth in the distribution system (Matilainen et al. 2010). However, it is important to note that the action of aluminum and ferric salts depends upon their hydrolyzed form and their soluble complexes that possess high positive charge as well as upon an acidic pH range.

Table 3 provides a summary of the finished water quality after the application of the combined pre-oxidation treatment for the raw water. Chen et al. (2009) reported that the presence of low calcium carbonate content (240 mg/L) in raw water was an important factor in the performance of the KMnO₄/alum system provided that the pH remained near alkaline. It was also reported that in the neutral pH range, surfaces of MnO₂ and algae cells developed a negative charge and calcium ions in water served to bridge and hold these negatively charged surfaces together into algal flocs (Chen et al. 2009). It can be seen from the data that ferric chloride and ferric sulfate provided slightly better finished water quality characteristics. However, the use of ferric chloride increased the final Cl as well as the use of alum and ferric sulfate increased the residual sulfates. Basiouny et al. (2008) studied the effect of KMnO₄ pre-oxidation alone upon the coagulation of raw Nile water. They found that an optimum KMnO₄ dose of 6 mg/L applied to Nile water resulted in a 35% turbidity removal. However, this required an alum dose of 40–60 mg/L and a ferric sulfate dose of 85–105 mg/L to achieve a turbidity reduction of 58% and 20%, respectively, with no pH adjustment.

Regarding the residual metal in final treated waters, accounts of the health effects and efficacy of using both Al and ferric salts in water treatment have been presented by Matilainen et al. (2010) and Simate et al. (2012). However, the use of alum as coagulant may lead to a high residual content as well, as it may form micro-flocs in the distribution networks which then provide nutrients for microbial growth (Lei et al. 2009; Matilainen et al. 2010; Sieliechi et al. 2010). On the other hand, it was reported that utilizing KMnO₄/alum in combination did reduce the amount of residual alum in the finished waters (Tomperi et al. 2013). Ferric salts are also favored as they are more robust operationally and insensitive to the coagulation pH (Wang et al. 2013). However, ferric salts were reported to induce high alkalinity consumption plus the residual sulfates or chloride in finished water may increase its corrosivity (Matilainen et al. 2010). High residual iron content in drinking water may also be responsible for disinfectant neutralization, discoloration and a metallic taste in water (Sieliechi et al. 2010). As ferric chloride is a corrosive compound that could rapidly attack metals, treatment plants using it as coagulant may require major modifications in dosage equipments and system (storage tanks, pumps, piping, valves and accessories) (Niquette et al. 2004). In addition, the use of ferric sulfate in drinking water treatment may incur higher operational costs than using alum; however, it offers a higher satisfaction to the consumer (Niquette et al. 2004).

The relationship between the Nile surface waters quality parameters and the occurrence of algae was investigated using multivariate statistical analysis. The data revealed that Chlorophyta and diatoms contents were the major contributors towards the measured total algal count and that the Euglenophyta content was strongly dependent upon the total alkalinity (CaCO₃) and the TDS of these waters. Cyanophyta also showed a strong dependency upon all chemical parameters measured as well as on Euglenophyta in these waters. As well, the effectivity of the current treatment procedure using pre-chlorination/alum on the clarification of the final treated water was investigated relative to the application of a combined KMnO₄/chlorine/coagulant stage. The results indicated that this application of KMnO₄ was effective in reducing the chlorine demand while maintaining an acceptable residual content. The use of the combined chlorine/permanganate pretreatment using alum, ferric chloride and ferric sulfate improved residual turbidity of the treated waters and enhanced percentage algal removal. Moreover, the finished water quality after the application of the combined pre-oxidation treatment for the raw water favored the application of ferric chloride and ferric sulfate. The study also revealed that the sulfates and chlorides ions contents were more responsible for the reduction in total algal count while the hydrolyzed metal ion was effective in the reduction of the final water turbidity. Economically, 1 kg of commercial permanganate costs 300 Egyptian pounds and 1 L of commercial NaOCl (14%) used at water treatment plants costs 30 Egyptian pounds. However, without changing much of the current treatment system, the application of permanganate would be safer and reduce the health hazards associated with the chlorine addition.