Performance and mechanism of ‘coagulation co-adsorption’ on simultaneous reduction of turbidity and low-fluoride pollution by a polymer metal-based defluorination and turbidity-removal agent (PMDTA)

Fluoride pollution in natural water bodies mainly results from natural and human factors (Lei 2022). The former mainly derives from the releasing of fluoride from ores containing fluoride under long-term natural factors, just like that of the releasing process of arsenic element (Bukke et al. 2024). And the latter mainly involves traditional and modern industries, such as inorganic material manufacturing (ceramics, cement, glass, brick, etc.), electric (thermal power generation, photovoltaic, etc.), pharmaceutical, chemical, electroplating, semi-conductor, metallurgical, and other industries (Jia et al. 2018; Sherly Williams et al. 2022; Neeti & Singh 2023; Awaleh et al. 2024). Therefore, fluoride pollution existing in the environment has been considered to be a serious public health hazard (Ahmad Dar & Kurella 2023) worldwide, mainly existing in natural waters and biological tissues. Fluoride pollution is widely distributed in Asia, Europe, America, and Africa (Suparna et al. 2021; Ahmad et al. 2022). More than 200 million people around the world drink water having excessive fluoride, based on statistics (Suparna et al. 2021), in which fluorosis cases caused by highly fluoridated drinking water were found to be distributed in at least 25 countries (Li et al. 2023).

Fluoride entered the food chain largely through water, soil, and air, but entered human bodies chiefly through drinking water (Lei 2022). Fluoride persists in the body for a long time once it enters, thus resulting in some serious health problems, such as dental fluorosis, skeletal fluorosis disease, arthritis, cancers, infertility, cardiovascular disease, thyroid disease, and Alzheimer's disease (Yadav et al. 2019; Han et al. 2021; Li et al. 2023; Takahiko et al. 2024).

China is one of the countries in Asia with the most serious fluoride pollution because it is located in the world's high fluoride belt (Podgorski & Berg 2022), for instance, fluoride reached up to 6.2 mg/L in the groundwater in Taiyuan Basin (Takahiko et al. 2024). There are more than 80 million people drinking water with excessive fluoride in China, especially the northern areas having relatively common endemic fluorosis (Jia et al. 2018), which mainly includes drinking water type, coal type, and tea type, among which drinking water is the most important type of endemic fluorosis (NHCPRC 2022). By the end of 2022, there were 1,042 areas (cities and districts) with endemic fluorosis (drinking water type) in China, including 248,000 dental fluorosis patients aged 8–12 years and 58,000 patients with skeletal fluorosis (NHCPRC 2022). In addition, fluoride is also toxic to animals and plants by having an effect on reproduction, growth, blood, and feeding efficiency of animals (Kazem & Salar 2023). Absorption of fluoride by plants (Yadav et al. 2018) causes physiological, biochemical, and structural damage, and even cell death.

Presently, the world's three famous international organizations and many other countries (more than 120) have put forward limits on fluoride level for drinking water quality standards (Li et al. 2023). For example, the fluoride limits are 1.5 (Fawell et al. 2013), ≤0.7 (EPA 2018) and 0.7–1.5 mg/L (Erika et al. 2005), respectively, according to the World Health Organization (WHO) Drinking Water Quality Standard, the United States Environment Protection Agency (EPA) Drinking Water Quality Standard, and the European Union Drinking Water Quality Standard (EC98/83). The limits for fluoride in drinking water in Japan and Thailand are 0.8 and 0.7 mg/L (MHLW 2020; Ahmad Dar & Kurella 2023), respectively.

China's drinking water standard (2006 Edition) also stipulates the limit of fluoride (1.0 mg/L), which is stricter than the WHO standard. The 2022 Edition of ‘Drinking Water Sanitation Standard’ (GB 5749-2022) still uses the 1.0 mg/L (SAPRC 2022) limit requirements. At the same time, the A degree discharge standard for wastewater containing fluoride in China is that the fluoride in effluent is lower than 10 mg/L (MEEPRC 1996).

Researchers around the world have been studying the technologies on removing fluoride for nearly 100 years (Khamkure et al. 2022; Liu et al. 2024). The common defluorination processes include chemical precipitation (You et al. 2023), ion exchange (Singh et al. 2020), membrane separation (Rathi et al. 2024), electrocoagulation (Castañeda et al. 2023), coagulation and precipitation (Jayasuriya & Nadarajah 2023; Kurniawan et al. 2023), and adsorption (Jeyaseelan et al. 2021; Huang et al. 2022). Each technology has its own limitations, even if it is a relatively common mature technology in removing fluoride, such as chemical precipitation, coagulation precipitation, and adsorption. Chemical precipitation is characterized by its simple operation, low cost, and larger water yield; however, it still has some negative effects, as follows: low efficiency, slow precipitation, and unstable quality of effluent, which is easily affected by raw water qualities and settling time and operating conditions (Lei 2022). Additionally, there are forms of secondary pollution, probably generating a large amount of dissolved aluminum and some complex sludge containing toxic fluoride and aluminum (Kumar et al. 2022). So chemical precipitation is generally used for treating highly fluoride waters (fluoride level greater than 1,000 mg/L) (Li et al. 2023), industrial wastewater, or can be combined with other technologies, such as coagulation, precipitation, and adsorption (Lei 2022; Meng et al. 2023). Coagulation–precipitation has the following advantages: easy operation, simple equipment, high removing efficiency of turbidity, low cost, large water production, and convenience in using for intermittent operation, but it has low defluorination efficiency when treating low-fluoride waters and could almost not reduce the fluoride to the limit required, which is less than 1 mg/L (Bukke et al. 2024). If aluminum coagulant is used, the effluent will contain a large amount of dissolved aluminum (Wang et al. 2020), which has lots of negative effect on human bodies and other animals or plants. The above problems limit the application of coagulation for removing fluoride to a certain extent in some occasions (Li et al. 2023). As for adsorption (N'Zébo et al. 2022; Huang et al. 2023), there are lots of researches on removing fluoride from different waters. Adsorption is characterized by its easy operation, low cost, good effect, basically no secondary pollution, and a wide range of adsorption materials. However, adsorption performance, which is less stable in removing fluoride, may be affected by many factors, such as adsorbent material, pH, fluoride concentration, other competing ions, contact time, temperature, and solid–liquid ratio (N'Zébo et al. 2022). Additionally, there are also other problems such as high regeneration cost, high personnel quality requirements, and inability to effectively remove turbidity and fluoride simultaneously (Takmil et al. 2020). Moreover, researchers have developed some advanced fluoride removal techniques, such as structural memory effect, ionic sieve effect, nano-surface effect, and anti-competitive adsorption in recent years (He et al. 2020). Of course, there are many combined processes, such as chemical precipitation + adsorption (Lei 2022), chemical precipitation + coagulation precipitation (Meng et al. 2023), and coagulation–precipitation + adsorption, but the combined processes have the problems of long process, poor coordination, and high cost.

Currently, many highly efficient fluoride removal technologies usually impose a significant economic burden on water or wastewater plants, placing heavy economic pressure on relatively backward regions (Zhang 2022). So, it is urgent to develop a fluoride removal technology that is both efficient and economical and easy to use (Zhang 2022), which can not only reduce the cost of fluoride removal, but also promote the popularization and implementation of environmental protection.

The key thought of this work is to embed the adsorption into the coagulation–precipitation process, called ‘coagulation co-adsorption (CcA)’, in which the adsorbent (that is the flocs) was produced by the coagulation process itself, no other additional adsorbents were required, and the flocs were directly treated and disposed of as the sludge. So, this work can not only reduce the treatment cost, but also directly reduce the fluoride level to lower than 1 mg/L, meeting the effluent standards without secondary treatment.

Traditional coagulants for removing fluoride mainly include aluminum-based and iron-based salt types. As is well known, the flocs generated by aluminum-based coagulants during the coagulation are smaller and denser, having higher coagulation fluoride removal than iron-based type. However, the flocs generated by iron-based coagulants have better precipitation properties. Therefore, the key focus of our work is to prepare an agent having the ability to remove fluoride and turbidity by combining the excellent defluorination performance of aluminum with the excellent precipitation performance of an iron-based coagulant. Poly-Si–Fe coagulant (PSF) is an excellent coagulant that was developed by our research group in the early stage (Fu et al. 2019), giving excellent turbidity removal and good floc precipitation performance (Fu et al. 2019). So, in this work, PSF was modified by an aluminum element to prepare a polymer metal-based defluorination and turbidity-removal agent (PMDTA) that can remove turbidity and low-fluoride pollution simultaneously. And then the CcA process was performed using PMDTA as the coagulant to form the PMDTA-flocs (their sizes were controlled during the coagulation process) as the adsorbent for adsorbing fluoride. At the same time, the reduction of turbidity was investigated simultaneously. Finally, the impact of precipitation and filtration on fluoride and turbidity removal was also studied using simulated fluoride water and real fluoride water samples, in comparison with that of PSF. It is expected to open up an important research direction for the realization of low cost, low environmental impact, and flexible operation of fluoride removal technology.

Liquid PSF was first prepared according to the literature (Fu et al. 2019). Aluminum (w(Al₂O₃) = 26%, purchased from Henan, China) was added to PSF at moderate stirring speed to make PMDTA with Si/Fe/Al molar ratio of 1/1/0.91. The PMDTA was then diluted to Fe concentration of 5 g (Fe + Al₂O₃)/L to use. The element composition of PMDTA was analyzed using the energy dispersive spectrum (EDS) method with GeminiSEM 300 Field Emission Scanning Electron Microscope (FE-SEM, Zeiss, Germany), in comparison with that of PSF, presented in Table 1.

The particle size of 50 times diluent of PMDTA and PSF were determined with Nano ZS90 Zetasizer (Malvern, USA).

The simulated and real water containing low fluoride were used as the test water samples whose qualities are summarized in Table 2.

The simulated fluoride water was made as follows: 0.12 g NaF and 400 mL soil leaching solution were added to 40 L tap water under stirring for 10 min to make fluoride concentration of 1.828–2.2 mg/L. The soil leaching solution was made as follows: 200 g soil derived from the campus at the University of Jinan was mixed with 2 L tap water under stirring for 30 min, followed by 10 min of settlement, and then was filtered with ordinary gauze to obtain the filtrate as the soil leaching solution, which was stirred for 1–2 min before use.

The real fluoride water was obtained from the effluent of primary treatment for the wastewater coming from a refrigerant chemical plant (Jinan, China), and its raw water qualities were as follows: Fluoride = 20–50 g/L, Turbidity = 0.663 NTU, Color = 0.261 A, = 2–3 g/L, SO^3– = 120–170 g/L, COD = 24–34 g/L, and pH = 13–14. The raw water was diluted by 10,000 times to obtain a diluent, and then 40 mL diluent was mixed with 40 L tap water and then mixed with 8 L of the soil leaching solution to obtain the test real water.

The essence of simultaneous removal of turbidity and low fluoride by the CcA method using PMDTA is to embed adsorption (to remove fluoride with flocs formed by PMDTA) into coagulation process (mainly to remove turbidity), achieving higher turbidity removal and also reducing fluoride to less than 1 mg/L for low-fluoride waters.

The jar test was used for CcA method, which was performed on the six-unit multiple stirrer system. First, the agent was added to the tested water (1 L) and rapidly mixed for 1 min at 400 r/min. The slow mixing stage of 30 min at 100 r/min followed to build up flocs having certain sizes that were beneficial to the adsorption of fluoride.

Sedimentation process was conducted with different times. Also, then the supernatant was withdrawn from a position of 2–3 cm below the surface for measurement.

The filtration test was performed on the self-made filter device (Fu et al. 2019) including influent tank, motor + mixer, peristaltic pump (BT100, Baoding, China), and filter column (diameter of 3 cm, height of 150 cm) with an overflow hole at 3 cm from the top to ensure that the filtration process is an equal-head filtration, rotameter (LZB, Changzhou, China), valves, and effluent tank. The sifted filter materials (quartz sand) and support materials (pebbles) were rinsed with deionized water to remove the impurities on the surface, and this was then followed by drying in GZX Drying Oven (Boxun, Shanghai) at 60 °C. The dried and cooled pebbles (thickness of 20 cm) and quartz sand (thickness of 80 cm and particle size from 0.5 to 1.5 mm with different size ranges for different purposes) were filled into the filter columns, accompanied by some gentle tapping to achieve a certain compactness for both the filter materials and supporting layers. The valve was first opened to the maximum flow in order to make the deionized water pass through the filter materials at full speed for 30 min, thus removing the bubbles among the filter materials. Then, the effluent coming from the treated water in Section 1.3.2 for precipitation of 60 min was passed into the filter column, and the initial filtration rate was set to 4 m/h.

The CcA test was the same as that in Section 1.3.1 and the acquisition of measured samples were as follows.

• (1) Without precipitation stage. After the test in Section 1.3.1, some mixture of the flocs and water without settle stage was withdrawn from jars for the following measurement.

• (2) With precipitation stage. Precipitation times of 15, 30, and 60 min were adopted after (1), respectively, and then the supernatant in Section 1.3.2 was withdrawn from the jars for the following measurement.

PSF was used as a compared agent, and the simulated and real fluoride water samples in Table 2 were used as the test waters. The dosage of PMDTA was selected from 40 to 90 mg/L, as Fe, compared with that of PSF. The samples from (1) and (2) (sedimentation time was 15, 30, and 60 min, respectively) in treating the simulated fluoride water and from (1) and (2) (sedimentation time was 60 min) in treating the real fluoride water were withdrawn for the measurement of residual fluoride (RF), residual turbidity (RT), and pH with fluoride ion detector with PF-2-01 fluoride ion electrode (Leici, Shanghai), PHS-3C pH meter (Mettler-Toledo, Switzerland), and HACH 2100 turbidity meter (HACH, USA), respectively. Three runs were performed in this test, the results represented the averages, and the error bars referred to the standard error of the mean of the three experiments.

The influence of other factors of water qualities on simultaneous removal of fluoride and turbidity by PMDTA were studied, such as pH, level of turbidity, fluoride, organic matters (humic acid (HA)), salt matters (as NaCl), and temperature. The simulated fluoride water in Table 2 was for the test in this section. The dosage of 60 mg/L was selected according to Section 1.4.1. The samples from (2) (sedimentation time was 60 min) for RF and RT measurement were the same as that in Section 1.4.1.

PSF was used as a comparing agent, and solid matters of PMDTA, PSF, and their flocs (formed after CcA tests) were made first: the liquid of PMDTA and PSF and their flocs were frozen under −4 °C for 24 h, and then were further frozen for drying under −99 °C in vacuum freeze dryer for 72 h to make their solids.

And then, the micro-properties of PMDTA were studied in comparison with that of PMDTA-flocs. The real fluoride water in Table 2 was for the test water in this section. The dosage of 50 mg/L was selected according to Section 1.4.1 apart from the following ‘Charged properties’ section.

After 30 s gold spray treatment, the surface morphology of solid PMDTA, PSF, PMDTA-flocsm and PSF-flocs were analyzed with GeminiSEM 300 Field Emission Scanning Electron Microscope (FE-SEM, Zeiss, Germany), respectively.

The bond structures of solid PMDTA, PSF, PMDTA-flocs, and PSF-flocs were studied with Nicolet iS50 Fourier infrared spectrometer (FTIR, Thermo Fisher, USA), respectively.

The Zeta potentials of the supernatant of the treated water by PMDTA at dosages of 40, 50, and 60 mg/L were studied with Nano ZS90 Zetasizer (Malvern, USA), and compared with those of PSF.

The test water for filtration was from the treated water for sedimentation of 60 min in Section 1.3.2.

• (1) CcA and sedimentation tests were conducted. A jar test was performed according to Section 1.3.1, in which PMDTA was used as the water agent, in comparison with PSF. The simulated and real fluoride water samples in Table 2 were used as the test waters. The dosages of 40–60 mg/L (different dosages for different test waters) were selected according to Section 1.4.1. The CcA procedure was the same as that in Section 1.3.1. After precipitation of 60 min, the RF and RT of the supernatant after CcA were measured as that in Section 1.4.1.

• (2) Filtration test as that in Section 1.3.3 was conducted. The RF and RT were measured (as that in Section 1.4.1) after filtration for 5, 10, 30, and 60 min, respectively.

The PMDTA dosage of 40 and 60 mg/L for the filter material size of 0.5–1.0 mm, and PMDTA dosage of 60 mg/L for the size of 0.5–1.0 mm and 1.0–1.5 mm were selected, respectively. The RT and RF were measured, respectively.

The dosages of 50 and 60 mg/L were selected, respectively. The RT and RF were measured, respectively.

In order to study the effect of PMDTA on simultaneous removal of turbidity and fluoride, PSF was used as the reference coagulant in this experiment, and the simulated and real fluoride water samples in Table 2 were used as the test waters. The impact of PMDTA dosage on simultaneous removal of turbidity and fluoride was investigated under different precipitation times, as shown in Figure 3. PMDTA-NS-S and PMDTA-S60-S referred to no settled stage and settled for 60 min in the simulated fluoride water with PMDTA, respectively, and so on.

For the simulated fluoride water, as seen in Figure 2(a), for PMDTA, the RF without precipitation was greater than 1 mg/L. After precipitation, the following relations were obtained. (1) The dosage had some impact on RF. The RF first decreased and then increased with the increasing of dosage. When the dosage was lower than 80 mg/L, the RF was lower than 1 mg/L, reduced to the lowest value of 0.78 at 60 mg/L (called the optimum dosage). (2) Precipitation time has only a little impact on RF. The RF curves for precipitation times of 15–60 min were basically overlapped. While for PSF, the RF was greater than 1 mg/L. Moreover, both dosage and precipitation time only had less impact on the removal of fluoride.

As seen in Figure 2(b), the removal of turbidity by PMDTA was slightly lower than that by PSF. For PMDTA, (1) the dosage had a slight impact on the RT, decreasing to 0.887 NTU at the dosage of 60 mg/L and precipitation of 60 min; (2) settling time had a great influence on the RT: almost lower than 4 NTU at settling time of 15 min. The RT decreased with the increasing of settling time, almost reducing to about 1 NTU after precipitation of 60 min. While for PSF, the RT basically decreased to lower than 3 NTU after settling for 15 min, and then further reduced to lower than 1 NTU after settling for 30 min.

Based on the results of the simulated fluoride water, only the supernatant samples after precipitation of 60 min of the real fluoride water were tested. As seen from Figure 2(a), for PMDTA, the RF at the other lower and higher dosages were all less than 1 mg/L apart from dosage 60 to 70 mg/L (RF was greater than 1 mg/L), especially reduced to 0.607 mg/L (the fluoride removal rate reached 51.22%) at the dosage 50 mg/L. While the RF of PSF was about 1 mg/L (the highest value was 1.1 mg/L), and the removal of fluoride was lower than 45%. As seen from Figure 2(b), the turbidity removal by PMDTA was higher than that by PSF. For PMDTA, the dosage had basically no effect on the RT, also indicating that the dosage in this experiment was slightly larger for removing turbidity, because the RT reduced to 0.26 NTU (the RT of PSF was 1.42 NTU at the same dosage) at dosage 40 mg/L. For PSF, the RT decreased with the increasing of dosage: it decreased from 1.42 to 0.711 NTU when the dosage increased from 40 to 90 mg/L.

Based on the removal of fluoride and turbidity, 60 and 50 mg/L can be used as the optimal dosage for MSPDF when treating the simulated and real fluoride waters, respectively, and 60 min can be used as the optimal precipitation time.

As displayed in Figure 2(c), for the simulated fluoride water, the acidity of the treated water by PMDTA was basically the same as that by PSF, which was closer to neutral. However, for the real fluoride water, the pH of the treated water by PMDTA was higher than that by PSF, the difference was about 2 units. The pH of the treated water by PMDTA and PSF was 5.83 and 3.63, respectively, at dosage of 50 mg/L and precipitation of 60 min. According to the acidity or alkalinity of the effluent, PMDTA was far superior to PSF.

In this section, the influences of pH, turbidity, fluoride concentration, HA concentration, NaCl concentration, and temperature on the RF and RT by PMDTA in treating the simulated fluoride water were analyzed, respectively, as presented in Table 3. Based on the experimental results in Section 2.2.1, PMDTA dosage and precipitation time were selected as 60 mg/L and 60 min, respectively.

As presented in Table 3, the neutral or weak basic water environment (pH = 7–8.5) was conducive to removing fluoride, in which the RF can be reduced to less than 1 mg/L (the removal rate was close to 50%). And lower or higher pH values (pH < 7 and pH > 8.5) had a negative impact on fluoride removal, which was lower than 20%, but the weak basic water environment was slightly superior to the acidic. In this work, the removal of fluoride mainly depended on its adsorption on the hydrolyzed flocs of PMDTA, while PMDTA was attributed to a free dilution process after being added to the test waters because it is a metal-based polymer, so is immediately hydrolyzed, thus leading to a great change in its form and surface properties like PSF (Fu & Gao 2011). Generally, the times required for the metal-based coagulant hydrolysis reaction and formation of metal complexes are very short, such as 3–10 s for Al(OH) ²⁺ and Fe(OH) ²⁺, in which 0.01 s was only needed for the formation of hydrated hydrogen–oxygen complex (François 1987). Therefore, after it was added to the test waters, PMDTA was hydrolyzed instantaneously to form a larger number of flocs (called PMDTA-flocs) first under some controlled conditions, carrying a wide range of charges (including positive charges), thus leading to an adsorption of fluoride on the PMDTA-flocs. Concurrently, electric neutralization occurred after the PMDTA-flocs migrated to the surface of the impurities and colloidal and adsorbed on them (François 1987). But the lower pH side was not conducive to form complexed hydrolysates charged positively, and the higher pH side was inclined to form some precipitates uncharged, so the above two conditions were not conducive to the adsorption of fluoride on the PMDTA-flocs that are not charged positively or uncharged, including electrical neutralization. In the whole test pH range, the RT was reduced to lower than 2 NTU, especially the higher pH side was more conducive to turbidity removal: RT reached 1.39, 1.17, and 1.24 NTU, respectively (corresponding to removal rates of 88.32, 90.17, and 89.58%, respectively) when the pH was 8.5, 10, and 11.5 (the turbidity removal was basically higher than 85%).

The different adapted pH for fluoride and turbidity removal by PMDTA further indicated that PMDTA gave different removing mechanism of fluoride from that of turbidity. Fluoride was mainly adsorbed on the PMDTA-flocs, in which the chemical adsorption by electrical neutralization was dominated. While turbidity removal by PMDTA was mainly based on a Derjaguin–Landau–Verwey–Overbeek (DLVO) theory which was proposed by Soviet scientists Deryaguin and Landau and Dutch scientists Verwey and Overbeek between 1940 and 1948 to explain the colloidal stability of the solution. The DLVO theory explains the interaction forces between charged colloidal particles, including van der Waals attraction and electrostatic repulsion caused by double layers. These two opposing forces determine the stability of the colloid. The DLVO theory is a basic theory for coagulation including double electric layer compression, adsorption/electric neutralization, adsorption/bridge, sweep/capture, and their synergic action. In other words, the removal mechanism of fluoride was relatively simple, while that of turbidity was complex, thus resulting in a smaller adapted pH range for removing fluoride (weak base or neutral) than that for turbidity (alkaline, neutral, and acidic). Table 3 indicates that the common favorable pH for fluoride and turbidity removal was around 8.5.

For the entire test turbidity range (initial range of 5.6–500 NTU), the RF was reduced to less than 1 mg/L (the removal rate was basically higher than 50%). Moreover, when the initial turbidity was less than 61 NTU, the fluoride removal increased with the increasing of the initial turbidity: RF decreased to 0.224 mg/L (at initial turbidity 61 NTU), and this was basically followed by a stable value of 0.3 mg/L with the further increasing of the initial turbidity. The RT decreased to less than 1 NTU (turbidity removal was more than 65%) for the entire test turbidity range. Table 3 indicates that an appropriate increasing in the initial turbidity was beneficial to fluoride removal, chiefly because more initial turbidity substances produced more flocs providing more adsorption sites for fluoride. Moreover, more turbidity substances itself also provided more adsorption sites for fluoride. Fluoride was first adsorbed on turbidity substances to form a combination of turbidity with fluoride, and then the combination substances together with PMDTA formed more flocs to achieve the increasing removal of fluoride. However, this also involved the related problems about increasing sludge treatment in the later stage, which was needed to be considered comprehensively.

The fluoride removal decreased with the increasing of fluoride level (initial fluoride) in the test water. The RF decreased to 0.084 mg/L only at initial fluoride of 2 mg/L. While at other initial fluoride quantities, although the fluoride removal tended to be stable or increased, the RF increased significantly. For example, when the initial fluoride was 100 mg/L, the fluoride removal reached 74.8%, but the RF was still up to 25.2 mg/L. This indicated that the fluoride removal by PMDTA with CcA method in this work can reach nearly 80%, which is probably the highest removal rate at this certain dosage. This was consistent with the above-mentioned mechanism of fluoride removal: the removal of fluoride by PMDTA was mainly through the fluoride adsorption on PMDTA-flocs, especially the dominant chemical adsorption; when the initial turbidity was a fixed value (only initial fluoride changed), the amount of flocs generated was basically within a fixed range, so almost providing a fixed range of surface area and adsorption sites for adsorbing a certain amount of fluoride, thus leading to a decreasing of fluoride removal with the increasing of initial fluoride. Within the entire initial fluoride range, PMDTA gave a relatively excellent removal of turbidity: the RT reduced to lower than 1 NTU (turbidity removal was more than 65%), indicating that the initial fluoride level had little influence on turbidity removal by PMDTA: the turbidity removal basically tended to be at a constant rate when the initial turbidity level was fixed (only initial fluoride changed in this test) (Table 3). This also suggested that PMDTA posed a very wide range of adapted fluoride if its purpose was to remove turbidity. However, in order to improve the simultaneous removal of fluoride and turbidity by PMDTA at a certain dosage, it was necessary to control the initial fluoride level within a certain range.

The organic matter (as HA) concentration in the test water (initial HA) had little influence on the removal of both fluoride and turbidity. When the initial HA was less than 200 mg/L, the RF reduced to less than 1 mg/L (decreased to the lowest 0.684 mg/L at HA 50 mg/L), and the RT was also decreased to less than 1 NTU. The RF and RT increased slightly with the further increasing of initial HA levels. This suggested that PMDTA had a very wide adapted range of organic matters when removing turbidity and fluoride simultaneously.

The initial NaCl concentration of the test water also had little influence on the removal of both fluoride and turbidity, of which the impact on turbidity removal was slightly greater than that on fluoride removal. For the entire initial NaCl levels, the RF reduced to lower than 0.9 mg/L (decreased to the lowest 0.617 mg/L at NaCl 5 mg/L), and the RT could also be reduced to lower than 0.8 NTU. With the largely increasing initial NaCl, the fluoride removal tended to be stable. For instance, when the initial NaCl was greater than 10 mg/L, the RF was basically stabilized at about 0.8 mg/L. This indicated that PMDTA also had a very wide adapted range of initial NaCl levels when removing turbidity and fluoride simultaneously, which has a larger practical significance.

As also presented in Table 3, the temperature of the test water (initial temperature) gave a certain influence on the removal of both fluoride and turbidity: both increased with the increasing of the initial temperature. When the initial temperature was lower than 20 °C, the fluoride removal decreased with the decreasing of the initial temperature (when the initial temperature was 5 °C, the RF decreased to 1.12 mg/L (the removal rate was only 38.4%)). This was because the lower temperature (especially below 10 °C) was not conducive to the rapid hydrolysis of PMDTA, resulting in a reduction of the surface area of its hydrolyzed substances and a further decreasing of adsorption sites, all decreasing the adsorption of fluoride on the PMDTA-flocs. However, when the initial temperature was up to more than 10 °C, the RF could be reduced to less than 1 mg/L. And when the initial temperature was above 20 °C, both RF and RT tended to be stable, indicating that PMDTA basically entered a stable phase of hydrolysis after reaching this temperature, and the properties of the hydrolytic substances also tended to be stable. In the entire initial temperature range, the RT could be reduced to lower than 1 NTU. So, PMDTA gave a relatively wide adapted temperature range of the water samples.

In a word, if considering the simultaneous removal of fluoride and turbidity, PMDTA had a very wide application range of initial turbidity, organic matter levels, NaCl levels, and temperature of the test water, in which the RF and RT could basically be reduced to less than 1 mg/L and 1 NTU, respectively.

Figure 3 shows the comparison of the surface morphology among PMDTA, PMDTA-flocs, PSF, and PSF-flocs (3 K× and 10 K×), respectively, when PMDTA and PSF were used in the real fluoride water experiments.

The essence of defluorination in this work was to control the excessive growth of the flocs formed during the flocculation process (controlling of energy input), in order to provide more adsorption space and sites of the PMDTA-flocs for adsorbing fluoride (of course, including the adsorption of other pollutants). Meanwhile, the adsorption time of fluoride on the PMDTA-flocs was controlled by controlling the stay time of the flocs floating in the test fluoride waters, in order to achieve a full fluoride adsorption, further reducing the RF to less than 1 mg/L.

As seen from Figure 3, the surface structure of PMDTA was very different from that of PSF. And the surface of the former was composed of relatively complex structures including planar, layered, porous, or slit-like and cotton-like clumpy structures (Figure 3(a) and 3(a1)), while the surface structure of PSF was relatively simple, mainly including a larger plane and clumpy structures composed of a concave–convex structure (Figure 3(c) and 3(c1)). Obviously, the surface area of PMDTA was larger than that of PSF, and the adsorption sites provided by PMDTA were also significantly larger than that by PSF. It could be thought that the surface structure of the flocs formed by PMDTA is complex and could provide more adsorption sites. The fluoride removal mainly depended on its adsorption on the PMDTA-flocs. So, this is one of the essential reasons why PMDTA gave higher fluoride removal than PSF (Figure 2(a)). However, the well-known DLVO theory was the main mechanism to remove turbidity, therefore, PMDTA was not superior to PSF in turbidity removal.

During the CcA test, PMDTA underwent a series of complex reactions such as hydrolysis and complexation. Additionally, the real fluoride water was used, so PMDTA hydrolysis products were also combined with other complex pollutants apart from combing with fluoride, forming PMDTA-flocs. The surface structure of PMDTA-flocs changed a lot after the CcA test, compared to that of PMDTA. The comparison between Figure 3(b)/3(b1) and 3(a)/3(a1) showed that the surface structure of PMDTA-flocs became more homogeneous compared with that of PMDTA, mainly consisting of planar, lamellar, and cotton-like structures, indicating PMDTA basically formed similar structures after combing with different pollutants. The PSF-Flocs' surface (Figure 3(d)/3(d1)) became quasi-three-dimensional, consisting mainly of flat, tightly connected chunks and cotton-like structures. So, it could be seen that the utilization of both space and adsorption site of the surface structure of PMDTA was significantly higher than that of PSF during the process of removing pollutants.

Second, the FTIR spectra of PMDTA (Figure 4(a)) was compared with that of PMDTA-flocs (Figure 4(a1)). Compared to that of PMDTA, all peaks of PMDTA-flocs became flatter and wider, demonstrating that PMDTA's structures became more complex and more obviously 3-dimensional after being combined with various contaminants in the real fluoride water. While the difference of the structures between PSF (Figure 4(b)) and PSF-flocs (Figure 4(b1)) was different from that of PMDTA (Figure 4(a) and 4(a1)), compared to that of PSF, the peaks became more irregular in PSF-flocs and many peaks presented to be sharper and narrower, in which there were some very narrow tiny peaks. This was consistent with the fact that the surface structures of PMDTA-flocs (Figure 3(b)) became relatively simple while the surface structures of PSF-flocs showed an enhanced stereotypic property, further demonstrating that the space and adsorption site utilization of the PMDTA surface structures were significantly higher than that of PSF and also indicating that PMDTA and PSF may have some essential differences in removing pollutants. For example, physical effect was dominant in the removal of fluoride and pollutants by PSF, but PMDTA posed much more chemical adsorption capacity, in which the chemical effect was higher than the physical, thus leading to a higher fluoride removal by PMDTA than by PSF.

Under the same background of the test waters and conditions, the different charges carried by PMDTA and PSF supernatant particles may be mainly because of the following two reasons. (1) PMDTA itself had a wider ZP distribution and a larger intensity range than PSF. (2) After being added to the real fluoride waters, PMDTA rapidly hydrolyzed to form the flocs (before being combined with pollutants) having a wider ZP distribution and positive charge distribution than PSF because PMDTA (containing Fe and Al) gave more metal elements than PSF (only containing Al). This resulted in that PMDTA residual particles in the supernatant gave different charge ranges after the CcA test. Obviously, based on the above two reasons, PMDTA gave different focus areas in removing pollutants by electric neutralization and adsorption compared with PSF. PMDTA will combine with pollutants having a wider range of charge, and will also combine with more negatively charged pollutants. Meanwhile, the combination between the PMDTA-flocs that were positively charged with the fluoride that were negatively charged was stronger than that of PSF. Therefore, PMDTA presented more advantages in removing negatively charged fluoride (Figure 2(a)). Combined with Figure 4, it is further proved that there may be some essential differences in the mechanism of fluoride removal between PMDTA and PSF, which was consistent with the above analysis.

The simulated and real fluoride waters in Table 2 were used in this section.

As seen from Figure 6(a), the filtration almost had little effect on the removal of fluoride. When filtration was performed for 5 and 10 min, the fluoride level was slightly reduced, but basically returned to the initial value of filtration with the increasing of filtration time. The above trend was basically the same for the different dosages of 40 and 60 mg/L. The influence of filtration on turbidity removal at lower dosage (40 mg/L) was greater than that at higher dosage (60 mg/L), because the filtration gave a strong ability to intercept the residual flocs in the coagulation supernatant, while the amount of residual flocs at 60 mg/L (the optimal dosage) was less than that at 40 mg/L. If the impact of filtration on the removal of fluoride and turbidity was considered comprehensively, it could be believed that the filtration only gave a limited ability to remove the pollutants having ionic properties, and also suggested the physical trap capacity may be much greater than the chemical capacity coming from the quartz sand, resulting in a lower removal of fluoride by filtration with quartz sand filter material and higher removal of turbidity substances. Moreover, there was a phenomenon of fluoride desorption in the filtration process when treating the simulated fluoride water, thus indicating that the chemical adsorption ability between the fluoride and quartz sand was relatively weak even if a little chemical interception occurred.

As shown from Figure 6(b), the impact of filter material size on fluoride removal was small. Moreover, for the filter material of 1–1.5 mm, there was a trend of desorption with the increasing of filtration time (such as filtration of 60 min), which was similar to the result of Figure 6(a). While for 0.5–1.0 mm filter material, there was no trend of fluoride desorption during the entire filtration time, although probably some desorption phenomenon will occur if the filtration time was extended further. As also seen in Figure 6(b), the influence trend of the two sizes on turbidity removal was basically the same: all reduced to the lowest value within 5 min filtration, and then basically remained stable with the increasing of filtration time. Therefore, if the impact of different filter material sizes on the removal of fluoride and turbidity was considered comprehensively, it could be believed that 1–1.5 mm was the better choice.

As seen from Figure 7, for both PMDTA and PSF, the filtration had a little influence on the fluoride removal: the fluoride concentration continued to slightly decrease with the increasing of filtration time. The fluoride levels were 0.72, 0.69, and 0.58 mg/L for filtration influent, filtration 10 min and 60 min, respectively, in comparison with 1.038, 0.958, and 0.92 mg/L for PSF, respectively. However, the fluoride level did not return to the initial value (the value of the filtration influent) with the increasing of filtration time, which was different from that of the simulated fluoride water. For turbidity, the influence of filtration on the RT at different dosages was basically the same, that is, all could be reduced to less than 0.1 NTU, especially reduced to 0.084 NTU at filtration of 5 min at a higher dosage of PMDTA (60 mg/L), and then maintained to a continuous slight reduction with the increasing of filtration time. While the RT of PSF was greatly affected by filtration, for instance, the turbidity of filtration influent was up to 1.5 NTU, and then basically reached the same as that of PMDTA after 5 min filtration (0.216 NTU), which also indicated that the ability of the quartz sand to intercept or adsorb the RT substances was much higher than that to fluoride.

Compared with the simulated fluoride water (Figure 6(a)), both CcA + precipitation and filtration had a relatively good impact on the removal of fluoride and turbidity in treating real fluoride water. For example, for the CcA process (corresponding to the influent entering into the filtration in Figure 6(a)), for the simulated and real fluoride waters at PMDTA dosage of 60 mg/L, the RF of the filtration influent was 0.957 and 0.782 mg/L, and the RT was 0.712 and 0.467 NTU, respectively. While after 5 min filtration, the RF was 0.919 and 0.72 mg/L, and the RT reduced to 0.182 and 0.084 NTU, respectively. This may be because the real fluoride water was composed of clay having a more complex composition, in which various substances were more closely connected or chelated due to physical or chemical forces, which was more conducive to their removal by coagulation/adsorption process.

After PMDTA was added to a low-fluoride water sample, lots of flocs having certain sizes were formed during CcA process with larger energy input (150 r/min for 10 min and 100 r/min for 30 min). During the entire CcA process, the mechanism of removing fluoride was mainly based on the following two types.

(1) One was the classic conventional DOLV flocculation mechanism. The negatively charged fluoride was electrically neutralized with the positively charged PMDTA hydrolyzates, and then adsorption/bridging occurred to form a large amount of flocs with certain sizes, further capturing the pollutants in the water samples by sweeping, thus leading to the removal of fluoride together with other pollutants. During this process, turbid substances were the primary nuclei together with PMDTA hydrolyzed substances for producing the flocs, thereby forming a large number of flocs with certain sizes.

(2) Another was the traditional adsorption mechanism. The adsorbents used in this work were the PMDTA-flocs having certain sizes and amounts produced during the coagulation process, in which both size and number of the flocs could be controlled. The purpose was to obtain the flocs with excellent adsorption performance for fluoride, which was mainly reflected from the larger surface area and more adsorption sites of the flocs. Chemical adsorption (higher than PSF, thus leading to higher fluoride removal than PSF) was dominant in this adsorption mechanism based on previous analysis. The coagulation flocs were complex complexation substances that were composed of the coagulant hydrolyzates and pollutants, so had quite complex surface structures and compositions. Therefore, there also existed physical adsorption (higher than PSF, thus leading to higher fluorine removal than PSF) apart from the chemical adsorption, as well as adsorption between the fluoride and other pollutants in a complex form.

After settlement, some residual turbid substances in the supernatant were removed further by filtration, but fluoride was only slightly affected by filtration.

• (1) For the simulated and real fluoride waters used in the work, the fluoride removal by PMDTA after CcA process was higher than that by PSF, in which PMDTA could reduce fluoride to less than 1 mg/L, reaching the lowest RF of 0.78 mg/L (at dosage of 60 mg/L) and 0.607 mg/L (at dosage of 50 mg/L) for the simulated and real fluoride water, respectively. While PSF decreased fluoride to higher than or equivalent to 1 mg/L or so, PMDTA aided in slightly lower turbidity removal than PSF in the simulated fluoride water, but had higher turbidity removal (reduced to 0.29 NTU at dosage 50 mg/L) in the real fluoride water than that by PSF (reduced to 0.687 NTU at dosage 50 mg/L).

PMDTA allowed higher fluoride removal at about neutral pH (pH = 7–8.5), and had RT lower than 2 NTU in the entire pH range of the test. PMDTA gave an adapted pH range of weak base or neutral for decreasing fluoride, but gave basic, neutral, and acidic pH range for reducing turbidity. The proper increasing of turbidity in raw waters was beneficial to removing fluoride, but the increasing of fluoride level had little influence on the removal of turbidity. PMDTA could be adapted to a wider turbidity, HA, and NaCl concentrations and temperatures of the raw test water when synchronously removing fluoride and turbidity: the RF and RT were basically all reduced to less than 1 mg/L and 1 NTU, respectively.

• (2) The surface structure of PMDTA was relatively complex (mainly consisting of plane, layer, porous, or slit-like and cotton-like clumpy structures), in comparison with a relatively simple surface structure of PSF (mainly including a larger plane and clumpy structures composed of concave–convex structure). PMDTA-flocs posed a larger surface area and more adsorption sites than PSF-flocs. Moreover, the former had significantly higher utilization of space and adsorption sites than the latter. They both had the similar bond structures. The ZP of the PMDTA supernatant at different dosages (40, 50, and 60 mg/L) had a wider distribution and coverage (−65 to 111 mV) than that of PSF ( − 65.2 to 9.87 mV) and the former also gave a wider distribution of positive charges (2.42–9.87 mV and 95.6–111 mV) than the latter (2.42–9.87 mV). The filtration had little influence on the fluoride removal, but had a certain degree of influence on turbidity removal. PMDTA was more suitable for practical application.

• (3) The mechanism of fluoride removal by PMDTA was relatively simple: mainly through its adsorption on PMDTA-flocs, in which electrically neutralized chemisorption was dominant. However, the turbidity removal mechanism was more complex, mainly based on the DLVO theory including double electric layer compression, adsorption/electric neutralization, adsorption/bridge, and sweep/capture.

PMDTA can efficiently decrease fluoride and turbidity to meet the treated water standard by utilizing the existing facilities in water plants.

• (4) Currently, we think the following three aspects can be further optimized to enhance fluoride removal by PMDTA's chemical adsorption.

First, the preparation process of PMDTA needs to be further optimized to obtain the surface structure of PMDTA for efficiently adsorbing fluoride, and to enhance the range and strength of the positive charges carried by the hydrolyzed flocs of PMDTA to improve its chemisorption capacity for fluoride.

Second, the CcA process conditions need to be further optimized to obtain the optimal size and surface structure of PMDTA-flocs having larger surface areas and more adsorption sites for adsorbing fluoride.

Finally, the filter materials need to be further studied to enhance the chemical interception and adsorption capacity for removing fluoride, so as to further enhance the removal of fluoride by the entire combined process of ‘CcA + sedimentation + filtration’.

The authors acknowledge the financial support of the Research Institute of Tibet Smart Water, University of Jinan (2024125). The authors also thank the editor and the anonymous reviewers for their valuable comments to improve the quality of this paper.

All relevant data are included in the paper or its Supplementary Information.

The authors declare there is no conflict.