Quality Assessment of Sludge from Filter Backwash Water in Swimming Pool Facilities

Abstract

Swimming pools are examples of water-intensive facilities, where solutions for reducing economic and environmental costs are searched with increased frequency. One of the solutions supporting savings is the recovery of water from wastewater, including backwash water obtained while rinsing the filter bed. The study objective was the qualitative and quantitative assessment of post-coagulation sludges, the main pollutant found in the washings. During the analyses, assessment of the sedimentation capabilities of the sludges was performed (gravitationally), particle size distribution was assessed (particle size distribution analyser) and assessment of phytotoxicity with the use of plant indicators in short-term tests was performed (Lemna minor, Lepidium sativum, Sinapis alba, Raphanus sativus). The samples were collected from two independent circulations, which differed in terms of capacity and type of coagulant used. The tested post-coagulation sludges were characterized by high content of total suspended solids: in samples from Circuit 1 from 251 to 128 mg/L, in Circuit 2 from 489 to 228 mg/L. However, the sedimentation processes enabled significant separation of sludges. The hydrolyzed coagulant contributed to the improvement of sedimentation capabilities of the sludges. Despite the fact that in many samples low sludge concentrations favored stimulation of plant growth, the post-coagulation sludges can constitute a hazard to plant growth, particularly in the long-term perspective.

1. Introduction

Considering the increasing frequency and duration of heat waves in Europe and the local shortage of freshwater resources, more attention should be paid to the possibilities of recovering usable water from waste streams [1,2]. The possibilities of utilizing washings from porous beds rinsing constitute a problem analyzed in the literature, e.g., heat recovery and use for flushing toilets [3,4,5]. The interest in water recovery and washings (backwash water) reuse will increase in the near future due to the regulations of the European Parliament introduced in 2020 regarding the minimum quality requirements for reused wastewater [6]. At the same time, washings from swimming pool facilities, due to its quality, offers much greater possibilities of recovery than wastewater [3,7].

The efficiency of the filtration process has a significant impact on the quality of the water in the pool [8]. Water in a closed circuit is purified and disinfected continuously and its losses are supplemented with feed water, e.g., tap water. In the process of water filtration, a gradual bed collimation takes place, which comprises the suspended solids, fibres and post-coagulation sediments attached to the grains. As a result, the space between them gradually fills up (bed porosity decreases) and the hydraulic resistance of the bed increases. In this process, rinsing water flows under pressure from the bottom upwards through the nozzles located at the bottom of the filter (in the direction opposite to normal filtration) [8,9,10,11]. Rinsing of pressure filters should last until the rinsing water (washings/backwash water) is completely clear. According to the recommendations, it is necessary to use from 4 to 6 m³ of water for each m² of filter bed for proper rinsing [12].

The operation of a swimming pool facility has a significant demand for water for domestic and household purposes. Apart from rinsing the filters, water for technological purposes is used to replenish losses in the circulation. Monthly water loss in a single pool is about 10% of its capacity (for a swimming pool of average capacity of 576 m³, it is over 57 m³). The daily recommended fresh water intake is 30 dm³ per swimmer [12]. However, in practice, this consumption varies and depends on the load and size of the facility (from 28 to 86 dm³/person) [13,14]. In addition, water is consumed for the sanitary needs of the facility users and staff and for cleaning works within the building and adjacent green areas.

For example, daily water consumption in the facility with a swimming pool, a leisure pool and a spa bath, in which water treatment circuits have a total capacity of 75.9 m³/h, is 9.88 m³/d (calculated based on the efficiency of the sample cycle). The volume of water consumed depends on the function and type of the pool, the method of technological solution, the efficiency of the equipment used, the attendance, the standard of equipment of the facility, the season of the year, the standard of living of the population and additional purposes [13,14].

The quality of the washings depends on many factors, including the length of the filtration cycle, the type and number of filters, the method of bed rinsing, the quality of supplementary water, the technology used, and the hydraulic conditions of the pool basin [3]. The backwash water are characterized by a large number of suspended solids and residues of coagulants, added to the treated water before entering the filter bed [3,15]. The concentration of organic matter in the washings is mainly concentrated around particles larger than 45 µm [15]. A high proportion of 30 µm fractions was also reported, as well as nanoparticles, which were approximately 955 nm in size [16].

Previous results of studies on evaluating the possibility of discharging backwash water into water or land show that it is necessary to apply sedimentation and dechlorination to reduce the most problematic physicochemical indicators [3,4,17]. The content of total suspended solids in the washings shows varied values ranging from 28 to 360 mg/dm³ [3,17]. They are characterized by high precipitation, ranging from 81 to 96% relative to their total volume. For example, a two-hour sedimentation process reduces total suspended solids from 360 to 84 mg/dm³ [17]. As a result of a 12 h free chlorine disappearance process, the value of this parameter can be reduced by up to 80% [3,17]. Sedimentation also allows for a partial reduction in COD [17].

Considering the significant share of sludge and suspended particles in the volume of backwash water, we should also analyse the potential possibilities of managing sludge from swimming pool washings—its volume share in the stream, its physicochemical properties, chemical stability and ecotoxicological risk. The management of waste sludge is an increasingly studied problem of the circular economy [18].

The aim of this study is a preliminary analysis of the physicochemical and ecotoxicological quality of gravity thickened sludge from washings collected in two pool water circuits.

2. Materials and Methods

2.1. Subject of the Study

The subject of the study were samples of sludge from backwash water collected after rinsing of the filter beds operating in swimming pool facilities. Figure 1 shows a block diagram of the presented research. The washings was collected in two municipal swimming pools once a week, in four independent samplings for each of the circuits.

In swimming pool no. 1, the samples were taken from the common circuit of the swimming pool and the slow lane (located by the slide)—hereinafter referred to as Circuit 1 (Circuit 1, the characteristics of the objects can also be found in the Supplementary Materials). Water in this circuit is purified by a multilayer sand bed with a hydroanthracite layer. There are three similar beds with 1800 mm diameter and filtration area of 2.54 m² each. The rinsing process is carried out manually with compressed air and water. The volume of water used for rinsing in this cycle is 4.75 m³/m² of the bed. Each of the beds is rinsed every 24 h, and the water is discharged into the municipal sewage system. The filtration process is accompanied by coagulation with aluminum sulfate (8.5%) and disinfection with sodium hypochlorite produced in situ by membrane electrolysis.

In swimming pool no. 2, the samples were taken from the common circuit of the swimming pool, the recreation pool and the slow lane, further designated as Circuit 2 (Circuit 2). Water in this circuit is purified on multilayer sand beds with a hydroanthracite layer, of various dimensions. In this circuit, there are two filters with diameter of 1800 mm (area of each bed is 2.54 m²), two filters with a diameter of 2350 mm (the area of each bed is 4.30 m²) and a filter with a diameter of 2000 mm (area of bed is 3.14 m²). Rinsing process is carried out with compressed air and water. The volume of water used for rinsing is 4.09, 4.16 and 4.39 m³/m² of the bed (given according to the order mentioned earlier). The beds serving the swimming and recreational pools are rinsed every 24 h, whereas the bed for the slow lane is rinsed every 48 h. The filtration process is accompanied by coagulation with aluminum hydroxychloride (10%) and disinfection with stabilized sodium hypochlorite.

Usually, in pool technology, the coagulant dose is selected based on the estimated load of the facility with bathers, circulation efficiency, and previous service experience. The dose may vary from hour to hour, resulting from manual adjustment. However, for the tested circuits, it is possible to determine the average dosing dose based on the efficiency of the circuit—the volume of treated water per unit of time. The capacity of Circuit 1 was 1500 m³/h, and the dosed average dose of aluminum sulfate (8.5%) was 0.75 g/m³. The capacity of Circuit 2 during the sludge sampling period was 4000 m³/h, and the dosed average dose of aluminum hydroxychloride (10%) was 0.54 g/m³.

The samples were collected into 10 L plastic canisters through drain channels during the rinsing process conducted after the facility was closed in the evening.

The sedimentation process in Imhoff funnels was carried out for 24 h; the change in selected physicochemical parameters (described in Section 2.2) was analysed during this time. After 24 h, the treated washings was decanted from the sludge. The sludge was collected for further analyses.

2.2. Characteristics of the Sludge

During sedimentation of the sludge, the changes in the content of total suspended solids and the increase in the sludge settled by gravity in the Imhoff funnel were measured. The total suspended solids (TSS) content was determined by the method of filtration through glass fiber filters (made by Hach with pore size 1.5 µm, sample volume 200 mL) [19]. The TSS results and sediment volumes [mL] presented are the mean value of four independent replicates. The graphs show the value of the arithmetic mean together with the standard deviation.

Then, samples of gravity thickened sediments were analysed with the use of a Mastersizer 3000 particle size distribution analyser (Malvern) with Hydro EV (dip-in wet sample dispersion), in the range of 0.01–3500 µm. The presented results are the average of five measurements taken automatically by the device.

Photographs were taken with an optical microscope of the samples of the tested sediments. At the same time, the assessment of the sediment phytotoxicity was started.

2.3. Assessment of the Phytotoxicity of Sludge

The gravitationally compacted sludge was subjected to ecotoxicological analysis. The influence of sludge in the samples was assessed for 10, 30, 50, 80, 100% of the hydrated sludge volume in deionized water. Control samples consisted of deionized water. The assessment of the changes in phytotoxicity of method based on the US EPA recommendations [20] using L. minor as the indicator organism. The assessment of matrices phytotoxicity was made based on the observation of either stimulation or inhibition of the growth in the number of fronds in a 7-day test (from day t₁ = 0 to day t₂ = 7). The average specific growth rate for a specific period is calculated as the logarithmic increase in the growth variables—frond numbers, using the formula below for each replicate of control and treatments:

μ_i-j =((ln(N_j) − ln (N_i)/t)

(1)

where: μ_i-j—average specific growth rate from time i to j; N_i—measurement variable in the test or control vessel at time i; N_j—measurement variable in the test or control vessel at time j; t—time period from i to j.

Percent inhibition of growth rate (I_r) may then be calculated for each test concentration (treatment group) according to the following formula:

%I_r = ((μ_C − μ_T)/μ_C) × 100

(2)

where: I_r—percent inhibition in average specific growth rate, μ_C—mean value for μ in the control; μ_T—mean value for μ in the treatment group.

The plant growth inhibition (phytotoxicity) indicator and growth inhibition coefficient values were positive, whereas the growth stimulation was indicated by negative values. The phytotoxicity tests of the samples were performed in parallel with the physicochemical assessment. To describe the results of ecotoxicological analyses, a unified concept of the toxic effect was used, denoted as E (%) [21]. Negative frond growth inhibition values mean stimulation of their growth. The samples are classified according to the magnitude of the toxic effect: I < 25%—non-toxic; I = 25.1–50%—low toxic; I = 50.1–75% toxic; I = 75.1–100%—highly toxic [21].

All samples in the inhibition test were carried out in triplicate, and the results were expressed as mean ± SD. The results presented graphically represent the mean value of all the independent trials performed.

Germination Inhibition Test

The assessment of the phytotoxicity of matrices using common radish (Raphanus sativus), watercress (Lepidium sativum) and white mustard (Sinapis alba) was made based on the Phytotoxkit^® test method [22]. In total, 5 mL of each of the test samples (10, 30, 50, 80, 100% of the hydrated sludge volume in deionized water) was poured onto Petri dishes, and then 10 R. sativus and S. alba seeds were sown on each of the samples, the plates were placed in a laboratory incubator (Elkon) at a temperature of 25 ± 0.5 °C. The number of sprouted seeds and the length of the roots were read after 72 h.

The presented results are the mean values of three performed repetitions. The phytotoxicity test included a test with fresh matrices as well as control. The examined effects of phytotoxicity included relative seed germination (RSG), plants relative root growth (RRG) and germination index (GI) [23]:

$$RSG(\%)=\frac{number of seeds germinated in test sample }{number of seeds germinated in control}\times 100$$

(3)

$$RRG(\%)=\frac{mean root lenght in test sample}{mean root lenght in control}\times 100$$

(4)

$$GI(\%)=\frac{RSG\times RRG}{100}$$

(5)

The toxicity classification is presented in

Table 1. Toxicity classification based on the germination index GI [23,24].

Germination Index Value	Effect
GI ≥ 100	Growth stimulation
100 > GI ≥ 80	Non-toxicity
80 > GI ≥ 50	Moderate toxicity
50 > GI	High toxicity

was used to interpret the germination index values [23,24].

All samples in the germination test were carried out in triplicate, and the results were expressed as mean ± SD. Means and standard deviation were calculated using the MS Excel statistical package. The results presented graphically represent the mean value of all the independent trials performed.

3. Results

3.1. Characteristic of the Sludge

The concentration of total suspended solids in the tested samples was high; however, it differed significantly depending on the individual sampling. The backwash waters collected from Circuit 1 contained TSS ranging between 251 ± 23 mg/L (Sample 1) and 128 ± 4 mg/L (Sample 3) (Figure 2a–d). Additionally, the highest value was recorded for Sample 1 (251.00 ± 23.48 mg/L), whereas the lowest was for Sample 3 (128.00 ± 4.76 mg/L). Sludge from Circuit 1 was characterized by susceptibility to sedimentation. Within the first 30 min, the content of total suspended solids was reduced to 130.25 ± 26.65 mg/L and 92.25 ± 8.66 mg/L, respectively, for Sample 1 and 3 (Figure 2a–d).

In this period, the increase in sludge on the bottom of Imhoff funnel was analysed by reading the volume. First readings after 30 min of sedimentation were: 5.2 ± 0.7 mL; 4.7 ± 0.6 mL; 4.0 ± 0.2 mL and 3.8 ± 0.3 mL, respectively, for Samples 1, 2, 3 and 4 from Circuit 1 (Figure 2a–d). With increasing sedimentation time, the volume of the sludge first increased and then decreased, and the sludge settled on the bottom of the funnel, subject to concentration. After 24 h of sedimentation, the volume of the concentrated sludge in the samples was: 4.9 ± 0.7 mL; 3.8 ± 0.8 mL; 4.1 ± 0.5 mL and 3.7 ± 0.3 mL, respectively, for Samples 1; 2; 3 and 4 (Circuit 1) (Figure 2a–d).

Prolongation of the sedimentation process to 24 h aimed at the removal of the highest amount of sludge possible and obtaining values compliant with the legal regulations [6]. This makes it possible to discharge the supernatant liquid into the environment. The majority of the tested samples met those requirements. In Circuit 1, only for Sample 4, the required removal of total suspended solids below 35 mg/L or 90% reduction (reduction level was 76.40%) was not obtained. The total suspended solids concentration after a day of sedimentation was: 27.75 ± 7.41 mg/L; 33.25 ± 3.77 mg/L; 25.25 ± 4.99 mg/L and 38.75 ± 2.99 mg/L, respectively, for Samples 1; 2; 3 and 4 (Circuit 1) (Figure 2a–d).

The analogous analysis was conducted for washing samples in Circuit 2. Washing samples from Circuit 2 were characterized by higher concentrations of total suspended solids. The TSS content ranged from 489 to 228 mg/L (Figure 2e–h). Additionally, the highest value was recorded for Sample 1 (489.00 ± 15.34 mg/L), whereas the lowest was for Sample 4 (228.00 ± 14.97 mg/L).

After 30 min of sedimentation, the total suspended solids concentration was reduced to the following values: 247.50 ± 11.73 mg/L; 259.75 ± 25.79 mg/L; 142.50 ± 10.85 mg/L and 133.25 ± 7.09 mg/L, respectively, for Samples 1; 2; 3 and 4 (Figure 2e–h). On the other hand, the TSS values after 24 h amounted to: 37.25 ± 6.65 mg/L; 35.00 ± 6.22 mg/L; 34.25 ± 8.66 mg/L and 35.25 ± 9.00 mg/L, respectively, for Samples 1, 2, 3 and 4.

Despite the higher content of total suspended solids in samples from Circuit 2 compared with Circuit 1, no significant increase in sludge volume could be observed. Mean volume of sludge in the samples after 30 min of sedimentation was: 5.2 ± 0.9 mL; 3.4 ± 0.5 mL; 3.8 ± 0.2 mL and 3.6 ± 0.32 mL, respectively, for Samples 1, 2, 3, and 4. After 24 h of sedimentation, the volume of sludge in these samples was: 2.4 ± 0.2 mL; 4.7 ± 0.2 mL; 3.8 ± 0.2 mL and 3.8 ± 0.29 mL (Figure 2e–h). It was also observed that Sample 1, Circuit 2 did not differ in terms of the course of the concentration process from other sludge samples analysed.

The concentrated sludge was compared in terms of the percentage contribution of a specified particle size in the concentrated sample. Distribution analysis demonstrated heterogeneity of the tested sludge samples (Figure 3). However, all samples were characterized by the presence of particles with diameters ranging between 1.0 and 1000 μm, and the maximum % of volume did not exceed 8%. In Sample 1: Circuit 1, the highest percentage contribution characterized particles in the range between 10 μm and 100 μm (Figure 3a). In the case of the Sample 2, the range of particle size was greater, and close to 4% of particles ranged between 100 and 500 µm (Figure 3b). In Sample 3, the distinction between two particle size ranges was even clearer, with over 4 % contribution of volume concerned their sizes between 40 and 50 µm and from 110 to 130 µm (Figure 3c). In Sample 4 Circuit 1 one particle size range was distinguished. Over 5% of volume contribution characterized samples from 10 to 80 µm (Figure 3d). The main methodological problem at this stage was the concentration of the sludge and the selection of measurement parameters so that the delicate flocs did not disintegrate (which can make reliable measurement difficult).

Samples collected from Circuit 2 demonstrated similar size ranges to the samples from Circuit 1 (Figure 3e–h). Sample 1 had the 10 ÷ 50 µm range, characterized by a close to 6% contribution in the sludge, and about 3% contribution characterized samples with size close to 110 µm (Figure 3e). Sample 2 exhibited over 6% contribution in the range between 20 and 50 µm (Figure 3f). Sample 3 was characterized by over 2% contribution in the range 60 ÷ 600 µm, and the particles in the range from 10 to 30 µm contributed to over 4% of the sample (Figure 3g). Sample 4 was characterized by the highest percentage share of samples in the range from 10 to 50 µm (Figure 3h).

Due to the possibility of diverse contaminants occurring in swimming pool backwash water, as well as in their sludge, the selected samples were subject to a microscopic observation. Apart from different sizes of sludge particles, typically forming aggregates, the samples were observed to feature: numerous fibres and fine fragments (below 10 mm) from swimwear (Figure 4a–d), as well as sand and hydroanthracite grains, which were flushed from the medium during the backflushing process (Figure 4a,c). Moreover, the sludge was observed to have small ciliates feeding on the flocculi, among others of the genus Colpidium and Paramecium (Figure 4d–h).

3.2. Phytotoxicity Assessment of the Sludges

Figure 5a–d present the toxic effect of sludges on the common duckweed Lemna minor—inhibition/stimulation of growth and the number of necroses observed (individual values are listed in Table S1 in the Supplementary Materials). It was demonstrated that samples of sludges differed in the level of the toxic effect depending on the circulation investigated (1 or 2), as well as the independent sampling (Sampling 1 ÷ 4). Moreover, the relationship between the increase in sludge contribution in the sample and the increase in inhibition was not noted in all of the presented cases. None of the samples collected in Circuit 1 (Figure 5a) demonstrated toxicity, mean values of growth inhibition did not show toxicity or were characterized by low toxicity. In the case of some samples, growth stimulation was observed (for instance, Sample 1 and 3 with a sludge contribution of 10%). In the samples with 100% sludge contribution, Samples 3, 4 exhibited low toxicity. The highest number of necroses was recorded in Sample 4, with 80% sludge contribution (Figure 5c)—45% of duckweed was damaged, and this result was reflected by the value of growth inhibition, which under such conditions was also at its highest—48.21% ± 10.20.

Samples from Circuit 2 were characterized by the higher inhibition of growth of Lemna minor (Figure 5b). All analysed samples with 100% sludge contribution and majority of samples with 80% sludge contribution were characterized by high toxicity. Significantly, Sample 2 was characterized by a high toxicity already at 30% sludge contribution, and in this case the toxicity was also reflected by the higher number of necroses found (Figure 5d), from 75 to up to 100% of duckweed coverage.

The second part of the phytotoxicity assessment focused on the parameters of root growth and germination rate—Lepidium sativum, Sinapis alba and Raphanus sativus (full data available in the Supplementary Materials Tables S2–S7). In tests with Lepidium sativum (Figure 6a–d) in the majority of samples analysed, the toxicity of the collected sludges was recorded (Circuit 1, Sample 1 in the range 30 ÷ 100%, Sample 2, sludge contribution: 30, 80, 100%, Sample 3 in the range from 50 to 100% of sludge). In the case of Sample 4 from Circuit 1 and Sample 3 from Circuit 2, toxicity was found throughout the range of sludge contribution (these data are also presented in the Supplement Table S4). On the other hand, Sample 2 from Circuit 2 was non-toxic/demonstrated growth stimulation. At the same time, in the majority of the analysed samples, the inhibition of germination was recorded, the value of which increased with the percentage contribution of sludge in the samples (Figure 6c,d). Higher toxicity was recorded for samples from Circuit 1, and thus the results were largely contrary to those recorded in the test with Lemna minor. Additionally, in tests with Sinapis alba (Figure 7a–d), the toxicity of the majority of samples from Figure 2 was obtained. These results were not confirmed in the test with Raphanus sativus (Figure 8a–d), in which only some samples with 100% post-coagulation sludge contribution demonstrated high toxicity (Figure 8a–d), whereas the majority of analysed sludges stimulated the growth of Raphanus sativus roots. Causes for this phenomenon should be searched not only in the diverse sensitivity of the test plant organisms to the investigated sludges, but also in the effect of pollutants present in the post-coagulation sludges on the growth and development of plants. Growth stimulation may be caused by low aluminum concentration [25,26,27].

Worth noting are the values of the RSG index (Relative Seed Germination, %), which are presented in the Supplementary Materials; in many samples, a relationship could be observed between the increase in sludge contribution in the sample and the inhibition of seed germination (Supplementary Materials: Tables S3, S5 and S7).

4. Discussion

The coagulation process is utilized in swimming pool facilities to support the filtration process by means of porous substrate, increasing the efficiency of removing organic compounds [28]. Coagulant is added directly upstream of the pump with prefilter, which ensures its good mixing with water. Subsequently, the precipitated sediment, depending on the size of aggregates is retained on the surface or gradually migrate inside the filtration medium which prolongs the contact time with pollutants in the water and increases the efficiency of their adsorption. Ready-to-use commercial products containing several percent aluminium salts (hydrolyzed or non-hydrolyzed) are typically used at swimming pool facilities [3,29].

In order to maintain good water quality in the pool it is necessary to regularly remove sediments and pollutants accumulated in the medium. The quality and number of sediments depend on a variety of factors, including the length of the filtration cycle, the load of the facility with people using the pool, hygiene habits of the pool users, frequency of exchanging water in the entire circulation, or the type and concentration of chemicals used for cleaning [3,17,30]. In the presented analyses, the quality of sediments clearly differed between the subsequent samplings, as well as between the circulations. Diverse sedimentation capabilities, as well as volumes of the sediment obtained, were characterized. Importantly, no clear relationship has been observed between the broader distribution of particles found in the sediment (their greater equivalent diameter) and their reduced settling velocity (Figure 2a–h). However, the results show that the quality of backwash water and the sediment present therein is difficult to estimate, and it often depends on the conditions that occurred in the given filtration cycle, e.g., sampling on a day of intensive use of the pool by school youth, swimming classes or days off work when the load on the facility is greater. Thus, the load is the key factor affecting the quality of washings and the number of pollutants deposited in the post-coagulation sludge.

The washings forming during the backwashing contains several percent of sediment (in the presented study this amounted to 2 to 7.5%). The analyses available in the literature focus on the sediments formed in the process of potable water treatment, where the sediment constitutes approximately 5% of backwash water volume [29,31]. Washings containing sediment is typically treated as wastewater and transferred directly to the sewage system. They include not only the precipitated flocculi, but also mechanical impurities (hair, fibres from clothing, the epidermis) and microorganisms, some of which are pathogenic [28]. Presence of this type of impurities was confirmed also for samples analysed by the author. The backwash water was found to contain bacteria Pseudomonas aeruginosa, Escherichia coli, coagulase positive staphylococci, Legionella sp. and Cryptosporidium Parvum oocysts [32,33,34].

Considering the chemical characteristics, the post-coagulation sediments can also contain minor amounts of metals, among others. Fe, Ca, Cr, Zn, Pb, Ni, Cu, Cd, Mg and Mn [28,35]. As well as considerable amounts of aluminium. Kluczka et. al., 2017, examined post-coagulation sediments (from the process of water treatment) comparing their quality with wastewater sediments, and determined that both sediments meet the requirements of the Polish Regulation of the Minister of Environment of 6 February 2015 on the concentration of heavy metals and can be used in agriculture. The total sum of lead, cadmium, nickel, copper and chromium was considerably lower than the permissible values, and the metals were found as residues (with the exception for cadmium bound with organic matter). Although the amount of tested exchangeable aluminium was low in both sediments, the concentration of the bioavailable aluminium in the post-coagulation sediment was considerable [35]. During the coagulation process it is highly important to maintain the neutral pH, which prevents the release of mobile aluminium (III) ions, which increase toxicity of the sediment and disturb the physicochemical properties of water. In practice, in swimming pool circulations the pH values range between 6.5 and 8.0 [36,37], which may increase the risk of toxic aluminium. Aluminium coagulants with differing characteristics were used in the presented circulations. In Circuit 1 it was non-hydrolyzed (aluminum sulfate (8.5%)), In Circuit 2 it was pre-hydrolyzed (aluminum hydroxychloride (10%)). Pre-hydrolyzed coagulants are characterized by the fact that they contain hydroxyl groups that determine their increased alkalinity. When polyaluminum chlorides are produced, sulfates and silica can be added to their solutions, increasing sedimentation of post-coagulation suspensions. It is estimated that lower doses of pre-hydrolyzed coagulants enable obtaining the same effects as in the case of non-hydrolyzed coagulants. Moreover, they reduce pH to a lesser degree, and their efficacy is not as strongly dependent on temperature [38]. In the analysed case, the type of coagulant was the determining factor for the sedimentation capabilities, but it did not determine the amount of sediment or particle size distribution. The determined mean TSS reduction level, the % was for Circuit 1, for Sample 1, 2, 3, 4, respectively: 88.96; 86.59; 80.27; 76.37, and for Circuit 2: 92.38; 92.65; 89.23, 84.54 (data not presented in Section 3). However, no direct relationship has been found between the type of coagulant (pre-hydrolyzed/non-hydrolyzed), and the toxic effect towards indicator organisms. The preliminary amount of total suspended solids did not impact the phytotoxic effect.

Despite the extensive presence of waste from water treatment processes, the literature on the topic of potential toxicity of post-coagulation sediments is rather limited and the results are often conflicting [39]. Traditional solutions in the field of waste management include the use of post-coagulation sediments as the agent for impurity removal (e.g., adsorbent of heavy metals and phosphorus) [29,40]. In the past, post-coagulation sediments were viewed as inert waste material, with low reuse possibilities, thus they were removed directly to waters. Only later were they considered toxic to living organisms due to the presence of aluminium [40]. That is why they are typically dried and stored or incinerated, which naturally brings about financial costs. Storage is linked to land use, and the incineration process is poorly accepted by the society and is efficient only in the case of sediments with low moisture content. That is why different ideas for the management of waste have been proposed, e.g., as an additive to construction materials [40,41,42].

The concerns related to the use of post-coagulation sediments in the contact with plants, in agriculture and gardening are mostly related to the potential negative effects of accumulation of certain heavy metals (in particular, aluminium), which may pose hazards to organisms of higher order. Moreover, the conditions ensuring a balance between aluminium and phosphorus ions are still not clear [43]. On the other hand, other authors have emphasized that due to the content of carbon, hummus substances and the alkaline soil, the addition of post-coagulation sediment may perform the role of a buffer [28,29]. The use of sediments may improve the structure, hydraulic conductivity, humidity and level of nutrients in soil, because it contains a considerable amount of macro- and micronutrients and organic matter [39]. However, development of modern analytics and toxicology brought a deeper insight in the negative effect of aluminium on the environment and living organisms.

The issue of aluminium migration from post-coagulation sediments is related to the acidic pH of soil, thus in the presented study an attempt was made to ensure neutral conditions (dissolving specific percentage contributions of sediment in deionized water). This will enable including a broader effect of environmental factors in future analyses. Phytotoxicity of aluminium is related directly to the environmental conditions that control solubility of the element in soil [39].

The ionic form of aluminium (Al³⁺) is believed to be toxic to plants already at micromole concentrations [44]. Some plants developed tolerance mechanisms through developing complexes of organic acids with aluminium, in the leaves or in the rhizosphere [44,45]. Being highly reactive, ionic aluminium hits the cell wall, cytoplasmic membrane, nucleus and the cytoskeleton of the plants’ roots. It affects the function of the mitochondria due to the overproduction of free radicals. Thus, it has a multi-level, harmful effect on plants. The symptoms also include morphological traits. Growth inhibition, reduced leaf surface area, wilting and increased incidence rate of chloroses are also observed [28,44]. These reports are confirmed in the study. The inhibition of growth, germination and increased incidence rate of necroses was often linked to the increased percentage contribution of sediments in the samples. However, the determining factor for the obtained level of toxic effect was the resistance of the plant indicator—the highest resistance to post-coagulation sediment in the case of Raphanus sativus, highest sensitivity in the case of Sinapis alba. Moreover, important differences in the values obtained between subsequent conditions were recorded, which confirms that the quality of backwash water and sludge changes with the subsequent filtration cycles.

Despite the knowledge of the toxic effect of aluminium on living organisms, possibilities of using post-coagulation sediments on agricultural areas are still considered. This stems mainly from the growing need to seek new solutions in the field of circular economy in the water and sewage sector. Maintaining the proper flow of raw materials—water and energy, while maintaining the natural flow of water, guarantees the regeneration of natural capital [45,46,47,48]. The concept focuses on using also waste raw materials, thus reducing their squandering. The management of water and sewage faces a multitude of challenges, because climate change is inherently linked to the intensification of the water crisis [18]. Water is the carrier of materials and energy, which ought to be used in a more sustainable manner in the modern world.

5. Conclusions

Preliminary assessment of the quality of post-coagulation sludges originating from rinsing of filtration media enabled formulating a series of observations and conclusions:

• The tested post-coagulation sludges were characterized by high content of total suspended solids.
• The sludges were concentrated gravitationally and the volume contribution of sludge in the backwash water was from 2 to 7.5%.
• The hydrolyzed coagulant (Circuit 2) contributed to the improvement of sedimentation capabilities of sludges.
• The concentrated sediments, apart from the flocculent suspension, contained numerous other solid impurities—material fibers, hair, sand and hydroanthracite particles (washed out of the filter bed during backwashing), as well as microorganisms. Sludge particles had a wide size distribution from 1.0 to 1000 μm.
• No direct relationship could be observed between the type of coagulant (hydrolyzed/non-hydrolyzed) and the toxic effect among the tested test organisms.
• The results of the toxicity assessment indicate that the post-coagulation sludge may pose a hazard to plants. Although growth stimulation was noted in some of the tested samples, as in the case of tests with Lemna minor and Raphanus sativus, it must be remembered that the consequences of long-term contact of plants with post-coagulation sediments from swimming pool facilities is unknown.
• In this case, the next research step is to extend the analyses with pot experiments, in which sludges could be an admixture to the soil.
• The presented results are also important from the point of view of the potential for the management of the backwash water itself. Recognizing the seriousness of the threats posed by post-coagulation sludge allows establishing a strategy for cleaning the washings before using it, for example, for greenery care.