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Currently, little is known about the impact of silver nanoparticles (AgNPs) on ecologically important microorganisms such as ammonia-oxidizing bacteria (AOB). We performed a multi-analytical approach to demonstrate the effects of uncapped nanosilver (uAgNP), capped nanosilver (cAgNP) and Ag2SO4 on the activities of the AOB: Nitrosomonas europaea, Nitrosospira multiformis and Nitrosococcus oceani, and the growth of Escherichia coli and Bacillus subtilis as model bacterial systems in relation to AgNP type and concentration. All Ag treatments caused significant inhibition to the nitrification potential rates (NPRs) of Nitrosomonas europaea (decreased from 34 to < 16.7 μM NH4+ oxidized day−1), Nitrosospira multiformis (decreased from 46 to < 24.8 μM NH4+ oxidized day−1) and Nitrosococcus oceani (decreased from 26 to < 18.4 μM NH4+ oxidized day−1). Escherichia coli-Ag interactions revealed that the percentage of damaged E. coli cells was 45% greater with Ag2SO4, 39% with cAgNPs and 33% with uAgNPs compared with controls. Generally, the inhibitory effect on AOB NPRs and E. coli/B. subtilis growth was in the following order Ag2SO4 > cAgNP > uAgNP. In conclusion, AgNPs (especially cAgNPs) and Ag2SO4 adversely affected AOB activities and thus have the potential to severely impact key microbially driven processes such as nitrification in the environment.
Engineered silver nanoparticles (AgNPs) are increasingly being used as a broad spectrum antimicrobial agent in a number of consumer products including medical equipment, clothing, pharmaceuticals, food storage containers, children's toys, cosmetics, optical devices and household appliances (Kim et al., 2007; PEN, 2013). However, AgNPs are readily released from these products during use, and so it is inevitable that they will enter the environment, with potential implications for key microbially driven processes (Moore, 2006; Nowack and Bucheli, 2007; Klaine et al., 2008). When AgNPs are released into natural systems, the chemical and physical properties of the nanoparticles may be transformed by biotic and abiotic processes, which in turn will influence their antibacterial activities. The presence of stabilizing or capping agents may influence the types of transformation that occur, for example, aggregation behaviour and dissolution rate (Tejamaya et al., 2012). In order to accurately assess the environmental risks of AgNPs, a better understanding of the effects of AgNP type and concentration on key functional groups of microorganisms is required. Despite this, knowledge of AgNPs in relation to ecologically important microorganisms is limited.
Nitrification is crucial in the global cycling of nitrogen and is a key step in the removal of nitrogen from environments such as wastewater treatment facilities (Prosser, 1989; Kowalchuck and Stephen, 2001). Ammonia-oxidation is the first and rate-limiting step of nitrification and is mediated by ammonia-oxidizing bacteria (AOB) and ammonia-oxidizing archaea (Prosser and Nicol, 2008). AOB are notoriously sensitive microorganisms, whose activity and viability are susceptible to changes in pH, temperature, salinity, light, dissolved oxygen and sulphide concentration (Joye and Hollibaugh, 1995; Strauss and Dodds, 1997; Rysgaard et al., 1999; Nicol et al., 2008). It is therefore not unreasonable to predict that AgNPs entering terrestrial and aquatic ecosystems (such as via wastewater treatment facilities) may disrupt ammonia oxidizers and hence nitrification rates.
The present study focused on two different types of AgNPs: uncapped (uAgNPs) and capped (cAgNPs), using different model bacterial systems to evaluate the environmental impact of such AgNPs on ammonia oxidizers, a key functional group of microorganisms involved in nitrification. We focused on methoxy-polyethylene-glycol (mPEG), a capping agent, which is commonly used to stabilize AgNPs and aid dispersion (Christian et al., 2008). Specifically, the study investigated the effect of different concentrations (0.5–50 mg l−1) of uAgNPs, cAgNPs and Ag2SO4 on nitrification potential rates (NPRs) using three different AOB species: Nitrosomonas europaea, Nitrosospira multiformis and Nitrosococcus oceani. The AgNP concentrations were selected based on the potential future environmental concentrations or concentrations that would arise in the event of an incidental spill (Gottschalk et al., 2009). The effects of uAgNPs, cAgNPs and Ag2SO4 on the growth and viability of Escherichia coli and Bacillus subtilis were also examined. The overall goal of the study was to use a multi-analytical approach and different model bacterial systems, to demonstrate the potential impacts of engineered AgNPs on a key functional group of microorganisms, important in nitrification in the environment.
Results and discussion
Characterization of uAgNPs and cAgNPs
The uAgNPs and cAgNPs were characterized in terms of their morphologies and size distribution (Table 1). The average diameter of the uAgNPs was found to be 118–188 nm, which was up to fivefold larger than the cAgNPs (average diameter 17–40 nm). Transmission electron microscopy (TEM) micrographs showed that the cAgNPs were predominantly present as single particles, while the uAgNPs tended to form aggregates (Fig. 1). Atomic force microscopy confirmed the cAgNPs were predominantly present as single particles (Supporting Information Fig. S1). Asymmetrical flow field flow fractionation confirmed that the cAgNPs had a smaller size distribution compared to the uAgNPs (Supporting Information Fig. S2). The data show that the average AgNP diameters varied depending on the technique used. This is not uncommon in nanoparticle studies and demonstrates the importance of using a multimethod approach for nanoparticle characterization (Domingos et al., 2009; MacCuspie et al., 2011; Baalousha and Lead, 2012).
Table 1. Characteristics of uAgNPs and cAgNPs in UHP water and LB broth
Nanoparticle type Medium Characteristic (unit) (method)
aZeta potential and Z-average diameter measurements are reported as the mean of triplicate runs, with five measurements made in each run.
bFlFFF was carried out on a single uAgNP sample and triplicate cAgNP samples.
cThe diameters of at least 100 cAgNPs and uAgNPs were determined from a minimum of three TEM micrographs.
dThe heights of 67 cAgNPs were determined from a minimum of five AFM micrographs.
eThe mean data for triplicate runs is presented.
Numbers given in brackets represent the standard error of the mean.
AFM, atomic force microscopy; BET, Brunauer, Emmett and Teller theory; DLS, dynamic light scattering; FlFFF, flow field flow fractionation; ND, not determined.
Surface area (m2 g−1) calculated from TEM average diameter
4.9 (± 0.6)
21.2 (± 0.1)
The ζ (zeta)-potential and aggregation behaviour of uAgNPs were both strongly influenced by the medium the uAgNPs were dispersed in. Specifically, the ζ-potential increased by almost twofold (from −32 to −17 mV) in Luria–Burtani (LB) broth compared with ultra-high purity (UHP) water (Table 1). In addition, the zeta average hydrodynamic diameter (measured by dynamic light scattering) increased by over twofold (from 188 to 495 nm) in LB broth compared with UHP water (Table 1). In contrast with the uAgNPs, the hydrodynamic diameter of the cAgNPs was unaffected in LB broth (35 nm) (Table 1). Dialysis measurements were performed to determine the rate of Ag+ dissolution from AgNPs and Ag2SO4. Results showed that there were no differences in the rate of Ag+ dissolution between the cAgNPs and uAgNPs measured over 51 h (i.e. encompassing the time the bacteria were exposed to AgNPs) (Supporting Information Fig. S3). However, the Ag2SO4 dissolution was generally greater than the nanoparticle dissolution over 51 h (Supporting Information Fig. S3).
Effect of AgNPs and Ag2SO4 on AOB NPRs
Nitrification, driven by ammonia oxidizers, is an important process in global nitrogen cycling. Herein, the effect of uAgNPs, cAgNPs and Ag2SO4 was investigated using pure cultures of AOB. It was notable that NPR inhibition varied depending on AOB species, AgNP type and AgNP concentration. All Ag treatments (uAgNPs, cAgNPs and Ag2SO4) significantly inhibited the NPRs of Nitrosomonas europaea (F18,56 = 153.7, P < 0.01), Nitrosospira multiformis (F18,56 = 177.6, P < 0.01) and Nitrosococcus oceani (F18,56 = 50.0, P < 0.01) (Fig. 2). In general, the inhibitory effect on AOB NPRs was in the following order: Ag2SO4 > cAgNP > uAgNP. Specifically, complete, or almost-complete, inhibition of Nitrosomonas europaea and Nitrosospira multiformis NPRs occurred when cells were incubated with Ag2SO4 between 0.5 and 50 mg l−1 (P < 0.01) (Fig. 2). While Nitrosococcus oceani NPRs were unaffected by 0.5 mg l−1 concentrations of Ag2SO4, at 5–50 mg l−1 Ag2SO4 almost-complete inhibition to NPRs occurred (P < 0.01) (Fig. 2).
When Nitrosomonas europaea and Nitrosospira multiformis were incubated with cAgNPs (at 0.5–50 mg l−1), NPRs were significantly inhibited (between 3.2 and 24.8 μM NH4+ oxidized day−1; P < 0.01) compared with controls without AgNPs (<46 μM NH4+ oxidized day−1) (Fig. 2). Interestingly, for Nitrosospira multiformis, NPR inhibition by 5 mg l−1 of cAgNPs was significantly less than that at 0.5 mg l−1 (P < 0.01). Also for Nitrosomonas europaea, NPR inhibition by 50 mg l−1 of cAgNPs was significantly less than that at 5 mg l−1 (P < 0.01) (Fig. 2). While 0.5 mg l−1 of cAgNPs did not affect Nitrosococcus oceani NPRs, cAgNPs (at 5–50 mg l−1) significantly inhibited NPRs (between 2.8 and 18.4 μM NH4+ oxidized day−1; P < 0.05) compared with controls (26 μM NH4+ oxidized day−1) (Fig. 1). The observed increase in NPRs from Nitrosomonas europaea and Nitrosospira multiformis grown with increasing concentrations of cAgNPs may be due to a dose-dependent effect on the behaviour of the nanoparticles (i.e. greater concentrations of nanoparticles lead to a greater likelihood of nanoparticle aggregation, which may result in reduced nanoparticle-AOB cell contact).
Nitrosomonas europaea NPRs were significantly inhibited with uAgNPs at 0.5–50 mg l−1 (3.1–6.3 μM NH4+ oxidized day−1; P < 0.01) compared with controls (34 μM NH4+ oxidized day−1) (Fig. 2). While Nitrosospira multiformis NPRs were unaffected by 0.5 mg l−1 of uAgNPs, 5 and 50 mg l−1 significantly inhibited NPRs (between 3.0 and 3.8 μM NH4+ oxidized day−1; P < 0.01) compared with controls (46 μM NH4+ oxidized day−1) (Fig. 2). Nitrosococcus oceani NPRs were significantly inhibited with 50 mg l−1 uAgNPs (2.8 μM NH4+ oxidized day−1; P < 0.01) compared with controls (26 μM NH4 oxidized day−1) (Fig. 2). Specifically, Nitrosomonas europaea NPRs were more severely affected by 0.5 mg l−1 uAgNPs than Nitrosospira multiformis which may be a result of physiological differences between bacterial species. Such findings have implications towards nitrification in wastewater treatment facilities where Nitrosomonas europaea are commonly found to dominate (Siripong and Rittmann, 2007).
Differences observed in NPR inhibition are also likely to be associated with the behaviour and aggregation state of the nanoparticles in the different growth media used. For example, Nitrosomonas europaea and Nitrosospira multiformis were grown in freshwater medium, whereas Nitrosococcus oceani was grown in seawater medium. It is likely that the greater ionic strength (3.5%) of the seawater medium caused the cAgNPs and uAgNPs to aggregate, resulting in the observed reduced inhibition by uAgNPs and cAgNPs of Nitrosococcus oceani NPRs (Christian et al., 2008; Handy et al., 2008b).
When Nitrosomonas europaea cells were exposed to mPEG (at equivalent concentrations found in 0.5–50 mg l−1 cAgNP), NPRs were significantly inhibited (between 2.5 and 23.7 μM NH4+ oxidized day−1; P < 0.01) compared with controls (34 μM NH4+ oxidized day−1) (Fig. 2). In addition, Nitrosospira multiformis NPRs were significantly inhibited when cells were exposed to mPEG, at equivalent concentrations found in 50 mg l−1 cAgNP (19.8 μM NH4+ oxidized day−1; P < 0.01), compared with controls (46 μM NH4+ oxidized day−1) (Fig. 2). It is possible that capping agents may also contribute towards nanoparticle toxicity (Handy et al., 2008b). In the present study, the mPEG capping agent did adversely affect the NPRs of Nitrosomonas europaea and (to a lesser extent) Nitrosospira multiformis. The exact mechanisms of mPEG toxicity to these AOB are not clear; however, we postulate that the mPEG may interact with the active site of ammonia monoxygenase (AMO), as it has been shown previously that organic compounds may directly bind to the copper atoms in the active site of AMO, inhibiting AMO catalytic ability (Bedard and Knowles, 1989).
Bacterial growth and viability when exposed to AgNPs and Ag2SO4
The potential antibacterial activity of uAgNPs, cAgNPs and Ag2SO4 was investigated using E. coli and B. subtilis (Fig. 3). Ag2SO4 caused the greatest inhibition of growth in both E. coli and B. subtilis, followed by cAgNPs and uAgNPs (Fig. 3). Although uAgNPs at 0.5 and 5 mg l−1 caused no significant difference in E. coli cell growth compared with controls, uAgNPs at 50 mg l−1 resulted in a significant reduction in cell density (1.1 × 106 cells ml−1; P < 0.05) by 12 h compared with controls (5.1 × 106 cells ml−1) (Fig. 3A). With 0.5 mg l−1 cAgNPs, E. coli cell densities were not significantly different from controls, but with 5 mg l−1 cAgNPs, cell densities were significantly reduced (0.6 × 106 cells ml−1; P < 0.05) after 12 h compared with controls (5.1 × 106 cells ml−1) and with 50 mg L−1 cAgNPs, E. coli cell growth was completely inhibited (Fig. 3B). Other studies have also shown that nanosilver particles are cytotoxic to E. coli (Pal et al., 2007; Yoon et al., 2007; Ruparelia et al., 2008; Martinez-Gutierrez et al., 2010). With 0.5 mg l−1 Ag2SO4, there was > 50% reduction to E. coli cell growth after 4 h until cell densities recovered to 3.2 × 106 cells ml−1 after 14 h, which was not significantly different from the controls (Fig. 3C). Furthermore, with 5 and 50 mg l−1 Ag2SO4, E. coli cell growth was completely inhibited (Fig. 3C).
With uAgNPs at 0.5 and 5 mg l−1, B. subtilis growth was not significantly different to controls (Fig. 3D). However, with 50 mg l−1 uAgNP, the final cell density of B. subtilis after 14 h was significantly reduced (7.9 × 106 cells ml−1; P < 0.005) (Fig. 3D). With cAgNPs at 0.5 and 5 mg l−1, B. subtilis cell growth was not significantly affected after 14 h, but with 50 mg l−1 cAgNP, cell growth was significantly inhibited, resulting in an extended lag phase and a final cell density of 1.2 × 102 cells ml−1 after 14 h (P < 0.001) (Fig. 3E). With 0.5 mg l−1 of Ag2SO4, the cell density of B. subtilis was not significantly different to controls, but with 5 mg l−1 Ag2SO4, there was an extended lag phase of 4 h, although cell densities after 14 h (6.1 × 107 cells ml−1) were not significantly reduced compared with controls (Fig. 3F). Ag2SO4 completely inhibited B. subtilis cell growth at 50 mg l−1 (Fig. 3F). Both E. coli and B. subtilis growth was not significantly affected by the mPEG capping agent alone, confirming that it did not contribute to the antimicrobial activity of cAgNPs (at cAgNP concentrations up to 50 mg l−1) (Fig. 3B and E).
In general, B. subtilis showed slightly more tolerance to nanosilver than E. coli, especially when exposed to Ag2SO4. In contrast with our study, others have shown that B. subtilis is more susceptible to nanosilver than E. coli (Yoon et al., 2007; Ruparelia et al., 2008). The differences in our results and that of others may be due to different strains used that demonstrate different nanosilver tolerances. In addition, in our study, the AgNPs were suspended in liquid broth compared with being placed on disks on agar plates as in the case of Ruparelia and colleagues (2008). One possible explanation for the differential toxic effects observed between E. coli and B. subtilis is that the AgNPs may bind to the teichoic acids found in Gram-positive cell walls and are in direct competition with divalent cations that may be present in the uncharacterized components of the yeast extract and tryptone in the LB broth, thus reducing nanoparticle bioavailability. However, studies by Chudasama and colleagues (2010), and Jin and colleagues (2010) have contradictory opinions of the role of the cell wall in reducing/enhancing nanoparticle-bacterial interactions and hence toxicity.
Specifically in our study, the cAgNPs were more toxic to both E. coli and B. subtilis compared with the uAgNPs. Herein, it was notable that the cAgNPs were less likely to aggregate both in UHP water and in LB broth, whereas the uAgNPs were present as aggregates in UHP water and formed larger particles in LB broth. Increases in nanoparticle size caused by aggregation can lead to a reduction in the specific surface area (dependent on size distribution and fractal dimension of aggregate) and thus influence nanoparticle toxicity (Handy et al., 2008a,b; Jin et al., 2010; Tejamaya et al., 2012). It is also possible that increased aggregation and therefore settling in the uAgNPs resulted in reduced exposure and thus lower toxicity. However, throughout the experiment, the flasks were gently shaken at 150 r.p.m. to reduce sedimentation of both the bacterial cells and AgNPs. It is therefore more likely that the observed reduction in the antimicrobial potential of the uAgNPs towards E. coli and B. subtilis when grown in LB broth was due to increased aggregation and thus reduced specific surface area of the AgNPs, which either reduced AgNP: bacteria interactions or reduced AgNP dissolution rates.
It is also possible that uncharacterized components of the yeast extract and tryptone in the LB broth (such as thiol-containing amino acids) influenced nanoparticle stability (i.e. the tendency of the particles to aggregate) and bioavailability. Furthermore, the NaCl content [i.e. 1% (w/v)] of the LB broth may have led to uAgNP instability. Even small increases in salinity can affect the surface charge of AgNPs, causing them to destabilize and aggregate (Christian et al., 2008; Handy et al., 2008a). In contrast with the uAgNPs, greater colloidal stability is likely to have been provided to the cAgNPs by the mPEG capping agent (Tejamaya et al., 2012).
To further investigate differences in antibacterial potential of the AgNPs, the Microtox assay was performed. Ag2SO4 was highly toxic with an effective concentration on 50% inhibition (EC50) of 1.73 mg l−1 (Supporting Information Table S1 and Fig. S4). Although EC50 values were not obtained for either uAgNPs or cAgNPs, EC10 values revealed 20% greater toxic response elicited by the cAgNPs (EC10 32 mg l−1) compared with uAgNPs (EC10 40 mg l−1). It is possible that the uAgNPs aggregated in the high ionic strength [2% (w/v)] Microtox diluent, which may have reduced its overall toxicity and would explain the differential toxicities observed between the cAgNPs and uAgNPs. Because the mPEG alone did not elicit a toxic response (at its equivalent concentration in up to 450 mg l−1 cAgNPs), it was assumed that the capping agent did not exacerbate the toxicity of the cAgNPs to Vibrio fischeri (Supporting Information Table S1 and Fig. S4).
Interactions between E. coli, AgNPs and Ag2SO4
To enhance our understanding of bacterial responses to specific AgNPs and hence possible causes of cellular toxicity, E. coli cells were incubated with uAgNPs, cAgNPs and Ag2SO4, and scanning TEM (STEM) performed to identify specific AgNP-bacterial interactions (Fig. 4). When E. coli cells were incubated with 50 mg l−1 Ag2SO4, there was a 45% increase in the number of damaged cells (either partially or completely) compared with controls. With cAgNPs, there was a 39% increase, and with uAgNPs, there was a 33% increase in the number of damaged cells (compared with controls). In some cases, where specific areas of cell membrane had been damaged (Fig. 4A, indicated by black arrows), E. coli cell contents were released into the surrounding environment. Energy dispersive X-ray (EDX) spectroscopy was applied to determine the specific location of Ag in and around E. coli cells, and although Ag was not directly detected within E. coli cells when incubated with uAgNPs, it was detected close to areas of damaged cell membranes in samples incubated with cAgNPs, and inside cell cytoplasm and around cell membranes with cultures incubated with Ag2SO4 (Fig. 4B, indicated by white arrows).
It is well known that AgNPs exhibit size-dependent toxicity, which may be due to either a direct effect of increased contact between the surface of the smaller particles and cell membranes (Sondi and Salopek-Sondi, 2004), or an indirect effect caused by increased dissolution of Ag+ from smaller AgNPs compared with larger AgNPs (Sotiriou et al., 2011; Ma et al., 2012; Xiu et al., 2012). It has been suggested that AgNPs may attach to and penetrate bacterial cell membranes (Sondi and Salopek-Sondi, 2004). Physical damage to cell membranes leading to increased membrane permeability, loss of membrane transport control and ultimately, cell death has been suggested as one of the primary causes of AgNP toxicity (Sondi and Salopek-Sondi, 2004; El Badawy et al., 2011). In the present study, increased bacterial cell membrane damage was observed in the presence of AgNPs and Ag2SO4. Furthermore, Ag was detected by EDX near to areas of E. coli cell membrane damage following incubation with cAgNPs, indicating a direct role for cAgNPs in bacterial cell membrane damage.
It is possible that an indirect effect of size-dependent AgNP toxicity was caused by increased dissolution of Ag+ from the smaller AgNPs. However, we have shown that there was no difference in the rate of Ag+ dissolution between the cAgNPs and the uAgNPs measured over 51 h (i.e. encompassing the time the model bacteria were exposed to AgNPs) (Supporting Information Fig. S3). Thus, size differences between cAgNPs and uAgNPs (including specific surface area and aggregation behaviour) rather than Ag+ dissolution are thought to have contributed to the observed differences in antibacterial activities herein. It is possible, however, that greater Ag+ dissolution from AgNPs may occur over longer periods of time (Kittler et al., 2010), which may have severe long-term implications to in situ environmental microbial communities and their processes. In the present study, Ag2SO4 (as a source of dissolved silver) had the greatest inhibitory effect on all of the bacteria tested. Ag+ dissolution from Ag2SO4 was greater than that observed for the AgNPs over 51 h (Supporting Information Fig. S3) and may account for the observed increase in toxicity compared with AgNPs. Several previous studies have also demonstrated that dissolved silver (usually in the form of AgNO3) is more toxic than AgNPs to bacteria (Choi et al., 2008; Jin et al., 2010; Pokhrel et al., 2012).
In conclusion, the inhibitory effect of the different Ag species was in the following order Ag2SO4 > cAgNP > uAgNP. It was notable that the cAgNPs had a greater inhibitory effect on AOB NPRs compared with uAgNPs. Generally, Nitrosomonas europaea demonstrated greater sensitivity to AgNPs (even at lower concentrations). Although the drivers for the observed differences in inhibition by the nanoparticles to the bacteria are not entirely clear, it is likely that size differences (rather than the capping agent itself) may have played a role (i.e. the cAgNPs were up to fivefold smaller than the uAgNPs). Our findings demonstrate that in systems such as wastewater treatment plants where nitrification-denitrifcation processes are important for nitrate removal, the accumulation of AgNPs is likely to have a detrimental effect on AOB activities (especially Nitrosomonas europaea), which may lead to system failures. Furthermore, following the release of AgNPs into receiving environments, the activities of AOB and faecal indicator organisms such as E. coli will be severely impacted. In nanoparticle studies, it is also important to investigate the effect of weathering processes on engineered nanoparticles (ENPs) and its effect on microorganisms, as exposure of organisms in the environment to ENP will occur not only via wastewater from manufacturers but also via municipal wastewater and direct inputs from weathering and ageing of products containing nanomaterials. It is therefore important that studies such as this focus on the impact of ENP on ecologically important groups of microorganisms like AOB.
This work was supported by a Natural Environment Research Council (NERC) CASE studentship with CEFAS (Ref: NE/H525289/1) and the University of Essex. The support of the NERC Facility for Environmental Nanoscience Analysis and Characterization is acknowledged. We thank Mr. Farid Benyahia for technical assistance and Dr. Terry McGenity for useful comments on the manuscript.