Loose deposits (LD) accumulate in drinking water distribution systems (DWDS) and may lead to tap water discoloration incidents upon resuspension. While inconvenient for the consumers and the water companies, discoloration may be accompanied by degradation of the microbiological quality of the water, and possibly to the transport of bacteria. The presence of heterotrophic bacteria towards LD with different characteristics was investigated. Bacterial loads in discoloured water samples and LD concentrated over different settling times were analysed. Total heterotrophic (TH) bacteria numbers did not relate to turbidity or to suspended solids in discoloured waters. Higher affinities of TH were observed for slower-settling LD (<24 h), which were also richer in microbial extracellular polymeric substances. The detection of viable iron-oxidizing bacteria in LD supported their possible roles in LD formation and may be related to microbial growth/regrowth in DWDS. Results suggest that LD may exhibit low affinities to hold and transport bacteria, at least to TH in chlorinated water supplies.

INTRODUCTION

Tap water discoloration is a frequent reason for customers to complain (Husband & Boxall 2011), particularly when recurrent. Discoloration arises from hydraulic perturbation and consequent resuspension (Vreeburg & Boxall 2007) of loose deposits (LD) that accumulate in drinking water distribution systems (DWDS). LD occurrence is generally associated with cast-iron corrosion (Benson et al. 2012), even if also occurring in non-metallic pipes (Vreeburg et al. 2008), and is often perceived solely as an aesthetic problem, owing to the yellow to brownish turbidity of discoloured water. However, as LD may adsorb substrates and nutrients, these may also be considered as prime sites for bacterial growth (Batté et al. 2003), including the hosting of pathogens, e.g., Legionella spp. (USEPA 2001). In fact, many problems in DWDS may be related to microbiological growth/regrowth, e.g., biofilm formation, LD accumulation, microbial influenced corrosion, opportunistic pathogens (Gauthier et al. 1999; Zacheus et al. 2001; Batté et al. 2003; Liu 2013). With respect to LD accumulation and further resuspension, discoloration incidents are often related to the microbial deterioration of drinking waters (Batté et al. 2003; Lehtola et al. 2004).

Possible sources for LD in DWDS include particles coming with upstream waters, or formed within the distribution network, e.g., through precipitation of bulk water colloidal and dissolved materials, pipe scaling and/or scouring, or biofilm detachment or sloughing (Gauthier et al. 1999; McNeill & Edwards 2001; Vreeburg & Boxall 2007; Husband & Boxall 2011). In addition, and apart from pipe corrosion, LD may enter the DWDS during pipe repair or rehabilitation procedures. Nonetheless, and despite so many LD origins, even if their formation mechanisms have not yet been comprehensively identified, discoloration LD typically have oxidized forms of iron and organic compounds (i.e., volatile solids (VS)) as main constituents (Gauthier et al. 1999; Barbeau et al. 2005; Vreeburg & Boxall 2007; Poças et al. 2013b).

Microbial extracellular polymeric substances (EPS) have also been reported as important LD constituents (Poças et al. 2013a). These, among a diversity of possible origins (e.g., raw water, pipe biofilms), may be produced by iron-oxidizing bacteria (FeOB) growing on appendages impregnated with iron oxides. These bacteria, apart from being associated with discoloration (Ridgway et al. 1981), may also produce organic substrates for microbial regrowth (Lehtola et al. 2004) during chemolithoautotrophic growth (Emerson et al. 2010). While embedded within the polymeric matrices of released or produced EPS' sheaths, micro-organisms may then grow protected against residual disinfectants or other oxidants (Flemming et al. 2007). Furthermore, EPS may contribute to the hydrogel floc nature of LD (Poças et al. 2013a), including to their high water contents and cohesive and flocculant properties (Vreeburg & Boxall 2007; Husband & Boxall 2010; Abe et al. 2012; Douterelo et al. 2012), and provide different sorption sites for bacterial development and growth. Thus, given that discoloration LD are hydrogel flocs (Poças et al. 2013a) with high interstitial water contents and permeability, discoloration LD may exhibit the ability to hold bacteria, which may differ from ‘true particles', such as sand grains or corrosion derivatives.

Although described as prime sites for bacterial regrowth in DWDS (Batté et al. 2003), the ability of LD leading to discoloration events to hold and transport bacteria is not clearly understood yet. With the purpose of evaluating the potential of discoloration LD for bacterial attachment and transport, the presence of total heterotrophic (TH) bacteria in LD with different characteristics (e.g., settling times, composition) was studied. The presence of viable FeOB in LD, due to their role in LD formation (Ridgway et al. 1981), and possible impacts in the bio-stability of DWDS water (Emerson et al. 2010), was also investigated.

MATERIALS AND METHODS

LD sampling

Samples were collected from the Lisbon DWDS. The water is mainly of surface origin (86%) and undergoes conventional treatment. A minor fraction of groundwater undergoes chlorination only. Samples were collected from pipes targeted by the water company cleaning programme, designed to respond to discoloured water complaints or reports. Sampling sites covered pipes from different and dispersed district metered areas, thus allowing for the sampling of pipes supplied with waters of different make-up and residence times.

Sampled locations (12) included dead-end and distribution pipes of different materials (asbestos cement (AC), ductile iron (DI) and high-density polyethylene (HDPE)) and ages (from 1 to 27 years). Cast-iron pipes, although present in the network (∼19%), were not part of the sampled LD sites. Of the 14 samples collected within the 12 locations, two were taken from the same pipes: one made of DI and the other of HDPE. Over about 2 months, large volume samples were collected in sterilized cylindrical containers (30 L) from each sampling site, along with unidirectional pipe routine discharges (Poças et al. 2013a, 2013b). Sampling velocities, rather than predefined, were those achieved by the hydrant's valve opening, thus allowing for selective and representative discoloration LD collection and characterization (Poças et al. 2013b).

LD separation

At the laboratory (Figure 1), 1 L aliquots of the large volume samples (W0h) were used for turbidity, conductivity, pH, total suspended solids (TSS) and TH enumeration analyses. For LD separation, the remaining volume of the large volume samples (c. 29 L) was left to settle at 4 °C for 24 h. Afterwards (Figure 1), 1 L of the top-supernatant water (SW24h) and 2 L of the bottom settled concentrates (LD24h) were collected with a peristaltic pump at a flow rate of 6 L/h. The SW24h and LD24h fractions were then analysed for TSS and TH. The remaining volume (c. 26 L) was left to decant at 4 °C (Poças et al. 2013a, b). After 7 days of settling (Figure 1), the supernatant volume was removed and c. 1–2 L of the enriched slurries (LD7days) was collected for LD physical–chemical characterization (Poças et al. 2013a, b). In this fraction, TH were not analysed, as bacterial proliferation is insignificant at 4 °C.

Figure 1

Sample separation: initial sampled waters (W0h), (SW24h and LD24h) fractions over 24 h of settling, and the concentrated slurries (LD7days), over 7 days of settling.

Figure 1

Sample separation: initial sampled waters (W0h), (SW24h and LD24h) fractions over 24 h of settling, and the concentrated slurries (LD7days), over 7 days of settling.

When performing laboratory tests, care was taken to reproduce the real system conditions as far as possible, including the LD physical features (e.g., shape, behaviour) and water matrix make-up.

Water and LD characterization

Turbidity (NTU) was measured in 25 mL borosilicate glass cells in a nephelometer (Turb 555–600100, WTW). Conductivity and pH were measured in a multi-parameter potentiometer (Consort, model C863). TSS were determined by the gravimetric method (APHA 1998). Total solids and VS were determined by drying at 110 °C to constant weight and ignition at 550 °C (APHA 1998). Total iron (Fet) was determined by the phenanthroline method (APHA 1998), with adaptations (Poças et al. 2013a, b). Polysaccharides (PL) and total protein (PT) were determined by the anthrone (Daniels et al. 1994) and Bradford (Bradford 1976) methods, as described in Poças et al. (2013a).

Microbiological analyses

In W0h, SW24h and LD24h, TH were determined by the most probable number (MPN) method in liquid R2A medium at 22 °C incubated for 7 days. The MPN method, which is appropriate for LD samples (Schaule et al. 1992), was not influenced by background turbidity, as it is distinguishable in colour and aspect from that due to TH growth. To preserve LD behaviour, and for comparison of results with previous publications (Zacheus et al. 2001), no sample pre-treatment was used, other than gravitational settling.

An adaptation of the gel-stabilized gradient tubes method (Emerson & Floyd 2005) was used to detect FeOB growth on Fe(II)-EDTA in O2 gradient cultures (Kumaraswamy et al. 2006). The presence of grown FeOB in the turbidity bands was further confirmed by microscopic observations of Prussian Blue (Pellegrin et al. 1999) stains. Bands' micro-organisms were also observed in fresh wet mounts under phase contrast microscopy and tested for Gram staining. Similarly treated smears of fresh Escherichia coli (ATCC 25922) cultures were used as controls. Tested inocula (n = 7 samples) were LD collected from a tap with discoloured water in a household with corroded DI plumbing and from the discharges collected from different DI (n = 4) and HDPE (n = 2) network pipes.

RESULTS AND DISCUSSION

Sampled waters

Large volume samples (n = 14) were collected during routine pipe cleaning at velocities from 0.07 to 0.32 m/s (Table 1) from the 12 sampling sites. Neither turbidity (from 1 to 46 NTU) nor TSS (from 0.3 to 35.8 mg/L) was influenced by the sampling velocities, the pipe materials, or the water characteristics, as shown by the variations in electrical conductivity (from 169 to 500 μs/cm) and in pH (from 7.5 to 9.9). This suggests that sampled waters were of different make-up and residence times.

Table 1

Sampling conditions and results observed for the 14 large volume samples

Pipe material Pipe age (years) Velocity (m/s) pH Conductivity (μs/cm) Turbidity (NTU) TSS (mg/L) TH counts/L  
AC 27 0.25 9.9 169 17 12.6 1.10 × 106  
DI 10 0.09 8.4 415 21 16.6 2.70 × 104  
DI 13 0.19 7.9 386 25 11.1 5.40 × 103 a(1st) 
DI 10 0.21 8.0 386 26 21.5 2.00 × 102  
DI 11 0.08 8.0 496 1.5 3.50 × 106  
DI 17 0.09 7.5 248 1.1 < 2.00 × 102  
DI 10 0.08 7.9 483 43 35.8 1.10 × 104  
DI 13 0.09 7.9 472 0.5 4.90 × 103 a(2nd) 
DI 14 0.09 7.9 500 2.0 4.50 × 102  
DI 0.07 7.5 298 4.2 < 2.00 × 102  
DI 12 0.32 8.8 253 0.3 2.10 × 106  
HDPE 0.18 7.8 438 9.0 9.20 × 105 a(1st) 
HDPE 0.18 7.8 373 46 19.0 5.40 × 106  
HDPE 0.20 7.9 392 27 21.0 < 2.00 × 102 a(2nd) 
Average 11 0.15 8.1 379 16 11.2 1.19 × 106  
St. dev. 0.08 0.6 102 16 10.6 1.80 × 106  
Max. 27 0.32 9.9 500 46 35.8 5.40 × 106  
Min. 0.07 7.5 169 0.3 2.00 × 102  
Pipe material Pipe age (years) Velocity (m/s) pH Conductivity (μs/cm) Turbidity (NTU) TSS (mg/L) TH counts/L  
AC 27 0.25 9.9 169 17 12.6 1.10 × 106  
DI 10 0.09 8.4 415 21 16.6 2.70 × 104  
DI 13 0.19 7.9 386 25 11.1 5.40 × 103 a(1st) 
DI 10 0.21 8.0 386 26 21.5 2.00 × 102  
DI 11 0.08 8.0 496 1.5 3.50 × 106  
DI 17 0.09 7.5 248 1.1 < 2.00 × 102  
DI 10 0.08 7.9 483 43 35.8 1.10 × 104  
DI 13 0.09 7.9 472 0.5 4.90 × 103 a(2nd) 
DI 14 0.09 7.9 500 2.0 4.50 × 102  
DI 0.07 7.5 298 4.2 < 2.00 × 102  
DI 12 0.32 8.8 253 0.3 2.10 × 106  
HDPE 0.18 7.8 438 9.0 9.20 × 105 a(1st) 
HDPE 0.18 7.8 373 46 19.0 5.40 × 106  
HDPE 0.20 7.9 392 27 21.0 < 2.00 × 102 a(2nd) 
Average 11 0.15 8.1 379 16 11.2 1.19 × 106  
St. dev. 0.08 0.6 102 16 10.6 1.80 × 106  
Max. 27 0.32 9.9 500 46 35.8 5.40 × 106  
Min. 0.07 7.5 169 0.3 2.00 × 102  

aRefers to the pipes sampled twice: a DI pipe at a 20 day interval and an HDPE pipe at a 49 day interval.

TH were detected in 11 out of 14 sampled waters at levels from 2.0 × 102 to 5.4 × 106 per L. Overall, sampled waters did not carry high TH loads, as these were within those typically found for drinking waters, where the average is around 5–10 × 106 per L (Zacheus et al. 2001; van der Kooij 2003; Allen et al. 2004), either with or without residual disinfectant.

TH loads were not related to sample turbidity or TSS levels (Figure 2). In fact, the highest loads (in the order of 106 per L) occurred in sampled waters with both low (<10 NTU) and high (>10 NTU) turbidity levels. In addition, at the locations where LD were sampled twice (Table 1), higher TH loads occurred in the less turbid samples, thus supporting the conclusion that TH loads were not related to sample turbidity or TSS levels.

These observations are consistent with those from previous studies where no clear relationship was found between turbidity and TH numbers (Liu 2013). On the other hand, a linear correlation (r2 = 0.82 and p-value <0.05) could be found between turbidity and TSS, which may suggest a similar LD behaviour in the collected samples. Therefore, and taking into account the observed different ranges of TH, LD bacterial content may have not varied due to different water turbidities (Figure 2), but possibly because LD were different in composition and/or age.

LD

Over the 24 h of settling, the average turbidity at the supernatant fraction was 6 ± 3 NTU, while that in the settled LD fraction was 91 ± 182 NTU. This shows there were still deposits settling after the first 24 h, which is consistent with previous studies where it is suggested that complete clarification of discoloured water samples takes place over several days (Poças et al. 2013b).

With respect to the bacterial loads in TH per gram, values found in SW24h and LD24h were compared to those measured in the collected water samples W0h (Figure 3). Relative to the changes in bacterial counts over the first 24 h, these should not have been significant, as the similarities on readings from the W0h and SW24h fractions show.

The two log difference in the unsettled (SW24h) and the settled (LD24h) LD fractions suggested that the TH loads were higher in the LD with the slowest settling rates, i.e., the SW24h fraction. Despite the fact that Figure 3 refers only to the averaged TH, that same trend was observed between W0h and SW24h, and, except in one case where TH were about four times higher in the LD24h fraction, between SW24h and LD24h. As TH per gram in SW24h and in W0h were also at comparable levels to those reported by Zacheus et al. (2001), which results were also obtained in R2A medium (≤109 TH/g of LD) and with the same sample treatment, it can be suggested that higher bacterial loads may be found in younger or in ‘softer’ LD (Zacheus et al. 2001). The different LD settling characteristics may have been due to dissimilarities in LD age and, likewise, in LD composition. This is in line with Liu (2013), who observed that the type of bacteria present in LD differed over accumulation and may be dependent on the amounts of LD used for characterization. Therefore, the differences in bacterial counts per mass of deposit may differ with LD settling rates and/or sampling sites, owing to different LD accumulation levels.

After the 7 days of settling, the settled LD fractions were used for physical–chemical characterization. Since only the slowest LD were left to settle after the first 24 h, LD characterization results mostly refer to the supernatant LD fraction (SW24h). On average, VS contents were 179 ± 103 mg/g and Fet contents 126 ± 56 mg/g (Table 2). This confirms that iron-rich LD may occur in all pipe types of the DWDS, given that sampled locations did not include pipes made of cast iron. Relative to the main constituents of EPS (Flemming & Wingender 2001), PT was only detected in four samples, while PL was found in all samples at considerable amounts (up to 268 mg/g). Although no correlations could be observed between LD components and TH (results not shown), these showed a potential relation with Fet and with EPS-PL components (Table 2), thus suggesting that LD composition may have had an influence on TH loads in LD.

Table 2

LD constituents and TH bacteria

Pipe material VS (mg/g) Fet (mg/g) PT (mg/g) PL (mg/g) TH/g 
AC 131 141 N.D. 135 6.2 × 108 
DI 157 83 N.D. 82 4.1 × 106 
DI 220 141 N.D. 105 2.8 × 105 
DI 133 119 N.D. 141 2.9 × 104 
DI 43 244 N.D. INT. 6.6 × 109 
DI 263 N.A. 268 N.D. 
DI 148 68 N.D. 56 5.9 × 104 
DI 125 N.A. N.D. INT. 2.4 × 106 
DI 168 93 INT. 2.5 × 105 
DI 67 73 N.D. 74 N.D. 
DI 471 N.A. 16 148 N.D. 
HDPE 250 N.A. 35 4.1 × 108 
HDPE 155 192 N.D. 152 1.8 × 109 
HDPE 171 107 N.D. 60 N.D. 
Average 179 126 114 9.4 × 108 
St. dev. 103 56 65 2.1 × 109 
Max. 471 244 16 268 6.6 × 109 
Min. 43 68 35 2.9 × 104 
Pipe material VS (mg/g) Fet (mg/g) PT (mg/g) PL (mg/g) TH/g 
AC 131 141 N.D. 135 6.2 × 108 
DI 157 83 N.D. 82 4.1 × 106 
DI 220 141 N.D. 105 2.8 × 105 
DI 133 119 N.D. 141 2.9 × 104 
DI 43 244 N.D. INT. 6.6 × 109 
DI 263 N.A. 268 N.D. 
DI 148 68 N.D. 56 5.9 × 104 
DI 125 N.A. N.D. INT. 2.4 × 106 
DI 168 93 INT. 2.5 × 105 
DI 67 73 N.D. 74 N.D. 
DI 471 N.A. 16 148 N.D. 
HDPE 250 N.A. 35 4.1 × 108 
HDPE 155 192 N.D. 152 1.8 × 109 
HDPE 171 107 N.D. 60 N.D. 
Average 179 126 114 9.4 × 108 
St. dev. 103 56 65 2.1 × 109 
Max. 471 244 16 268 6.6 × 109 
Min. 43 68 35 2.9 × 104 

N.D. = not detected; N.A. = not analysed; INT. = interferences.

The different LD settling rates suggested that slower-settling LD could be closer to soft pipeline young deposits, which is in agreement with what had already been suggested by Zacheus et al. (2001). Likewise, slower LD may be closer to hydrogel flocs, i.e., flocs with fibrillar matrices and densities approaching water (Poças et al. 2013a), apart from having a fractal three-dimensional structure. These hydrogel flocs, unlike heavier ‘true particles' that can settle more rapidly (e.g., pipe and encrustation derivatives, sand particles), may be easily transported throughout the network upon resuspension and reach consumers' taps during discoloration events. Therefore, to analyse the potential health risks occurring upon LD accumulation, not only are the microbial identification and quantification required, but also the probability of sampled LD reaching the consumers' taps needs to be evaluated. Thus, as in water safety plan methodologies (Bartram et al. 2009), both the microbial hazards and the probability of LD reaching the consumers' taps need to be evaluated when assessing LD potential risks that may occur upon LD resuspension.

Iron bacteria

The occurrence of Gram-negative motile rods with the ability for microaerophilic growth coupled to Fe(II) oxidation at circumneutral pH was observed in the analysed LD samples from DI pipes (n = 5). This suggests a possible involvement of FeOB in discoloration. Likewise, since chemolithoautotrophic Galionella spp. may be associated with discoloration LD (Ridgway et al. 1981; Li et al. 2010), carbon fixation products may be formed within the LD matrices, then allowing for increases in the overall AOC budget and, consequently, to bacterial regrowth. As detected FeOB grow under low oxygen concentrations (Emerson et al. 2010) it is also possible that, similarly to within the pipe-wall biofilm EPS matrix, other micro-organisms, including anaerobes, survive within the LD floc microenvironments (Liu 2013), where they are protected against residual disinfectants. The assessment of the potential health risks occurring during LD resuspension and tap water discoloration may need, therefore, to be complemented with the enumeration of specific bacterial species, other than TH only.

Bacterial affinity towards LD with different characteristics

In this study, TH loads were low in the sampled waters (<107 TH/L) and did not correlate with discoloration intensity (Figure 2). This is consistent with previous studies on LD, where TH were not associated with turbidity or TSS, possibly due to particle-associated bacteria and, thus, to TH underestimations (Liu 2013). Likewise, TH numbers may only refer to a small fraction (<0.1%) of the total quantity of microbes present, thus adding difficulty to the finding of correlations.

Overall, these results, although obtained from a single water supply system, which is also chlorinated, contrast with descriptions of LD as prime sites for bacterial growth (Zacheus et al. 2001; Batté et al. 2003; Liu 2013), either from chlorinated or non-chlorinated systems. This may be due to the different water characteristics, the analytical methods, the sample treatment (Liu 2013), or the levels of residual disinfectants (Camper 2004). On the other hand, investigations into the microbiological quality of drinking waters usually disregard the different accumulation times of sampled LD. In addition, overestimations of LD potential health hazards may be led by the collection of non-discoloration representative LD. Comparative studies on three-stage samples (i.e., biofilms, LD and the bulk water) should also address possible differences within sample residence times and behaviour (e.g., LD with different settling rates), as well as describe differences in the sampling procedures in use (e.g., pipe velocities and sampled volumes).

As Table 1 shows, the finding of trends between the water characteristics and turbidity or TSS may be difficult to achieve, possibly due to the differences in the water and the LD residence times. However, while LD behaviour and sampling velocities may be quantified, sample residence times are more difficult to assess. As hypothesized herein, lighter LD may refer to the younger LD fraction. If confirmed, the relation between LD behaviour and age could help to estimate sample residence times and, therefore, contribute to a better understanding of the risks associated with tap water discoloration. In terms of DWDS' operational and maintenance procedures, indications could also be provided on how to assess the potential health hazards of discoloration LD samples with different settling rates, which, in turn, may be an indication of the LD resuspension potentials.

CONCLUSIONS

The presence of TH in LD from DWDS was investigated. TH loads in discoloured water samples and in LD that were concentrated over different settling times were analysed. Higher affinities were observed for slower-settling LD, which were also richer in microbial EPS. Overall, results showed that the intensity of discoloration was not related to high TH loads in a chlorinated DWDS. Despite this, different bacterial loads and types of bacteria may be expected in LD with different characteristics (e.g., settling times and behaviour, composition or residence times) and collected using different sampling methods. Furthermore, the detection of chemolithoautotrophic iron-oxidizing bacteria in LD, apart from the role of FeOB in discoloration, suggests a possible association with bacterial regrowth in DWDS.

On the whole, the assessment of microbial health hazards of LD accumulation should rely not only on microbial identification and quantification data, but also on sample residence times and LD behaviour (e.g., settling and resuspension), as well as on LD sampling procedures (e.g., pipe velocities, sampled volumes). Also, as lighter LD may correspond to younger LD, better knowledge on the association of LD behaviour with age (e.g., through LD settling/resuspension velocities and EPS contents) could help the assessment of the potential health risks arising with tap water discoloration on consumers. Similarly, it could provide indications of the necessary pipe cleaning frequencies, thus supporting water companies in designing LD control programmes.

ACKNOWLEDGEMENTS

This work was funded by the Portuguese Foundation for Science and Technology (FCT) PTDC/ECM/108261/2008 and SFRH/BD/43715/2008 project and doctoral grants, respectively. The authors are also grateful to Joana Aguilar, Ana Ribeiro, Joaquim Rosário and José Osório (EPAL) for their valuable technical support.

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