Doolette et al. Chemistry Central Journal 2013, 7:46
http://journal.chemistrycentral.com/content/7/1/46

RESEARCH ARTICLE

Open AccessTransformation of PVP coated silver nanoparticles
in a simulated wastewater treatment process andthe effect on microbial communities
Casey L Doolette1*, Mike J McLaughlin1,2, Jason K Kirby2, Damien J Batstone3, Hugh H Harris4, Huoqing Ge3
and Geert Cornelis5

Abstract
Background: Manufactured silver nanoparticles (AgNPs) are one of the most commonly used nanomaterials in
consumer goods and consequently their concentrations in wastewater and hence wastewater treatment plants are
predicted to increase. We investigated the fate of AgNPs in sludge that was subjected to aerobic and anaerobic
treatment and the impact of AgNPs on microbial processes and communities. The initial identification of AgNPs in
sludge was carried out using transmission electron microscopy (TEM) with energy dispersive X-ray (EDX) analysis.
The solid phase speciation of silver in sludge and wastewater influent was then examined using X-ray absorption
spectroscopy (XAS). The effects of transformed AgNPs (mainly Ag-S phases) on nitrification, wastewater microbial
populations and, for the first time, methanogenesis was investigated.
Results: Sequencing batch reactor experiments and anaerobic batch tests, both demonstrated that nitrification rate andmethane production were not affected by the addition of AgNPs [at 2.5 mg Ag L-1 (4.9 g L-1 total suspended solids, TSS)
and 183.6 mg Ag kg -1 (2.9 g kg-1 total solids, TS), respectively].
The low toxicity is most likely due to AgNP sulfidation. XAS analysis showed that sulfur bonded Ag was the dominant
Ag species in both aerobic (activated sludge) and anaerobic sludge. In AgNP and AgNO3 spiked aerobic sludge, metallic
Ag was detected (~15%). However, after anaerobic digestion, Ag(0) was not detected by XAS analysis. Dominant
wastewater microbial populations were not affected by AgNPs as determined by DNA extraction and pyrotag
sequencing. However, there was a shift in niche populations in both aerobic and anaerobic sludge, with a shift in AgNP
treated sludge compared with controls. This is the first time that the impact of transformed AgNPs (mainly Ag-S phases)on anaerobic digestion has been reported.
Conclusions: Silver NPs were transformed to Ag-S phases during activated sludge treatment (prior to anaerobic
digestion). Transformed AgNPs, at predicted future Ag wastewater concentrations, did not affect nitrification or
methanogenesis. Consequently, AgNPs are very unlikely to affect the efficient functioning of wastewater treatment
plants. However, AgNPs may negatively affect sub-dominant wastewater microbial communities.
Keywords: Silver nanoparticles, Silver sulfide, Wastewater treatment, STEM HAADF, Sequencing batch reactor,
Nitrification, Microbial communities, Pyrotag sequencing, Silver speciation, XAS, Synchrotron, Biosolids* Correspondence: casey.doolette@adelaide.edu.au
1
School of Agriculture Food & Wine, The University of Adelaide, PMB 1, Glen
Osmond, SA 5064, Australia
Full list of author information is available at the end of the article© 2013 Doolette et al.; licensee Chemistry Central Ltd. This is an Open Access article distributed under the terms of theCreative Commons Attribution License (http://creativecommons.org/licenses/by/2.0), which permits unrestricted use,
distribution, and reproduction in any medium, provided the original work is properly cited.

Doolette et al. Chemistry Central Journal 2013, 7:46
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Background
Rapid expansion of the nanotechnology industry has
occurred over the previous decade. Manufactured
nanomaterials (MNMs) encompass a variety of engineeredmaterials, which can be divided into two groups for thesake of clarity: nano-sized particles (having at least two
dimensions < 100 nm) and secondly, materials that are notparticulate but have nano-sized properties [1] (i.e. enhanced
electronic, optical and chemical properties compared to thebulk material). Silver (Ag0) nanoparticles (NPs) are themost widely used NPs in both consumer products and in
medical applications [2]. The anti-bacterial properties that
render AgNPs desirable may lead to increased risks tohuman and environmental health following release into theenvironment. The primary exposure pathway of AgNPs
into the environment is via wastewater streams. Silver
NPs may enter wastewater through the washing of Ag
nano-containing textiles [3,4] or plastics [5], or as aresult of the use of nano-enhanced outdoor paints [6]and washing machines [7].
Several authors have investigated the fate of manufactured
AgNPs in wastewater treatment plants (WWTPs) and havereported that the majority (> 85%) of AgNPs will be captured by biosolids (stabilised sludge) [5,8-11]. Accordingly,
the predicted effluent concentrations of AgNP are very low(ng L-1) [11], whereas AgNP concentrations in sludge are
predicted to be much higher (1 – 6 mg Ag kg-1) [11]. Bothconcentrations are likely to increase as the AgNP producing
industry expands. Given this scenario, and the strong
anti-bacterial effects of AgNPs, the stages of WWT thatare likely to be affected by AgNPs are those that are
dependent on the efficient functioning of microbes.
Such stages are the aerobic activated sludge process andanaerobic digestion, which proceeds the former process
in most WWTPs. There are very few studies that haveinvestigated the impact of AgNPs on both processes in asequential manner. Given that the transformation of
AgNPs is likely during WWT [8,12], it is crucial tounderstand at what stage transformation occurs so
accurate risk assesments can be conducted using AgNPs
in realistic forms.
During the activated sludge process, organic nitrogenand phosphorus are removed by various microbialcommunities. Several studies have investigated the impact
of AgNPs on nitrification [9,13-15] and the effects on
microbial populations that perform these processes[16,17]. However, results from nitrification studies aredivergent with no inhibition [9] and varying degrees ofinhibition [14,15,18] observed on nitrification following
AgNP addition in WWTPs or bioreactors at concentrations
between 0.4 and 1 mg Ag L-1.
The observed variation is most likely explained by thedifferences in input variables. A number of parameters
differ between studies, all of which are known to influence

Page 2 of 18

AgNP fate and toxicity e.g. intrinsic AgNP properties(size, coating), Ag concentration, sludge/wastewater properties (temperature, ionic strength (IS)), total suspended
solids (TSS) and dissolved organic carbon, (DOC)), thetype of sludge/wastewater used (realistic or artificial)and general experimental set-up (e.g. light intensity andwavelength which may cause photocatalytic reductionof Ag+ and AgNP).
The impact of AgNPs on anaerobic digestion has beenless studied than that of nitrification. Methanogenic
microorganisms are generally less sensitive to toxicants
than aerobic communities. Silver NPs have been shownto have no effect on biogas and methane production atconcentrations of 40 and 85 mg Ag L-1, [19] and [13],respectively.
The bactericidal mechanism of AgNPs (and Ag+) to organisms is only partially understood and debate is ongoingas to the exact means of action [20]. However, there isconcern that the same properties that render AgNPs usefulas an antimicrobial may also impact WWTP performanceby affecting sludge microbial populations. A high diversityof bacterial populations in WWTPs is crucial for successful
removal of BOD/COD, SS and biological phosphorus
and nitrogen.
The effects of AgNPs on sludge microbial communities
have been investigated by a limited number of studies. The
model nitrifying bacteria Nitrosomonas europae has beenshown to be adversely affected by AgNPs at concentrationsof 0.3 mg Ag L-1 [16] and 2 mg Ag L-1 [17]. These aremuch higher Ag concentrations than would normally befound in the environment at present. The microbialcommunities found in anaerobic systems generally havea different response to toxicants compared to aerobic
communities and are usually more sensitive to surface
active and homeostatic inhibitors and less sensitive tometabolic inhibitors [21] . For example, at very high Ag
concentrations (40 mg L-1), methanogenic communities(Methanosaeta and Methanomicrobiales) have beenshown to be largely unaffected by AgNP exposure [19].
So far, however, there has only been analysis of dominantmicrobes, through relatively insensitive techniques such
as qPCR, without assessing the impact on subdominant
populations as allowed by next generation techniquessuch as t-RFLP.
This study was undertaken to (i) investigate the effects
of Ag and polyvinylpyrrolidone coated (PVP) AgNPs onorganic nitrogen removal from wastewater (nitrification)(ii) examine the fate of Ag+ and AgNPs during various
stages of WWT (iii) investigate the effects of transformed
Ag+ and AgNP on anaerobic digestion efficiency, and (iv)
to determine if dominant and niche microbial community
structures in aerobic and anaerobic sludge are impacted by exposure to transformed Ag+ and AgNPs
using pyrosequencing.

Doolette et al. Chemistry Central Journal 2013, 7:46
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Results and discussion
Silver nanoparticle partitioning in the sequence batch
reactor process

Measured concentrations of Ag as a function of time in
the mixed liquor and effluent are shown in Figure 1.
Silver concentrations in the mixed liquor of eachsequencing batch reactor (SBR) increased non-linearly
during the 10 d aerobic stage. The cumulative concentration of Ag in the mixed liquor was less than the nominal
value (taking into account Ag losses with effluent)possibly due to losses of mixed liquor that occurred
during sampling for nitrification analysis and during
decanting. In addition, Ag losses may have been due
to sorption/complexation of Ag/AgNPs onto SBR tubing
and container walls.
The effluent concentrations of Ag in the SBR spiked
with AgNPs varied from 0.5 μg L-1 (day 5) to 7.7 μg L-1(day 1). This corresponds to between 0.1% (for days 5 –8)
and 5.4% (day 1) of the total amount of Ag in the mixed
liquor (nominal) being removed with the effluent. Similarly,
in the Ag+ dosed SBR, between 0.1% (days 5 – 8) and 1.1%
(day 1) of Ag was released with the effluent. Surprisingly,
the Ag concentration range of the effluents collected fromthe control SBR [0.3 μg L-1 (days 2 and 7) to 4.7 μg L-1(day 8)] were within the same range as the effluentscollected from the Ag+ dosed SBR; 0.3 μg L-1 (days 7 and 8)
to 4.7 μg L-1 (day 4). This can be explained by backgroundAg concentrations in the influent wastewater (15.0 ± 7.6 μg
Ag L-1) and activity sludge mixed liquor. Overall, theaverage (n = 8 days, where day 1 and 8 are the first andlast days of Ag addition) percentages of Ag in theeffluents ± standard deviation (SD) were 0.8 ± 0.1%, 0.4 ±
0.4%, and 2.0 ± 2.8%, for the AgNP, Ag+ and control SBRs,
respectively. There is large variation in the control as the
background Ag concentrations were close to inductively
coupled plasma-mass spectrometry (ICP-MS) instrumental
detection limits (0.05 μg/L). The results demonstrate thatthe majority of Ag spiked into SBRs, as AgNPs or ionic Ag+
was retained by the sludge.
The partioning results in this study are in agreementwith previous studies which have shown that the majorityof AgNPs in wastewater will be partitioned to the sludge
fraction following wastewater treatment [8-10]. However,the degree to which AgNPs are removed from wastewaterhas varied between each study. In a pilot WWTP experiment [8], 2.5% of spiked Ag (added as AgNPs stabilisedby polyoxyethylene fatty acid ester) was released from the
WWTP with the effluent, whereas in a 15 d simulated
SBR experiment (0.9 L working volume), citrate coated
AgNPs were found to be completely removed from thewastewater [9]. In the literature, the lowest removals of
AgNPs from wastewater (88 ± 4%) were recorded from a
SBR experiment using synthetic wastewater and AgNPs
with an unspecified polymer coating [10].

Page 3 of 18

The observed variations in removal efficiency of
AgNPs from the above studies may be due to a numberof factors including; the intrinsic properties of the NP(i.e. size, surface charge and capping agent) which in turnare influenced by additional parameters (e.g. mixed liquor
pH, chloride concentrations, etc.), method of spiking [16]and perhaps most importantly, the characteristics of theinfluent wastewater and activated sludge. The TSS content
of the influent and activated sludge determines the initial
mixed liquor TSS. In the current study, the TSS content(4.5 ± 0.6 g TSS L-1) was greater than that used in otherstudies [8-10](3, 2.4 and 1.8 g TSS L-1, respectively). This
may explain the high removal efficiency of AgNPs fromwastewater (> 99%) which we observed in this study. Most
NPs in WWTP sludge is likely to be heteroaggregatedwith bacteria [10,22-24] but NPs can also be associatedwith iron oxides or other inorganic particles [23].
At the conclusion of the SBR experiment, sludge Ag
concentrations were 418, 168 and 6 mg Ag kg-1(TS) for
the AgNP, Ag+ and control treatments, respectively. The
high concentration of Ag in the AgNP sludge is dueto the higher Ag concentrations in the AgNP spiking
suspensions (mean ± SD; 39 ± 6 mg Ag L-1, n = 11)
compared to the Ag+ solution (mean ± SD; 19.4 ± 0.1 mg
Ag L-1, n = 3); rather than a greater removal of Ag from thewastewater fraction. Due to logistical limitations, theconcentrations of each AgNP suspension could not be
determined before spiking as the homogenised NP
suspensions degrade after 24 h.
The Ag concentrations of the prepared AgNP suspensions were higher than we had previously achieved andtherefore higher than the nominal spiking concentration.
The total mass of Ag added to the AgNP and Ag+ SBRs,
not including background inputs from effluent, was12.7 mg and 6.4 mg, respectively. The results from the
SBR experiments show that the majority of AgNP is
partitioned to the solid phase.
Silver nanoparticle transformation during the sequencebatch reactor process as determined using STEM analysis

Numerous bright regions were observed in sludge collectedfrom the AgNP dosed SBR (Figure 2a) using scanningtransmission electron microscopy (STEM) analysis in highangle annular dark field (HAADF) mode. Further analysisof the bright spots by energy dispersive X-ray analysis(EDX) confirmed that these regions contained Ag (Figure 3).
The STEM image shows aggregates of Ag approximately
100 – 120 nm in diameter (Figure 2b). The higher magnification image (Figure 2b.) shows that each aggregate appears
to consist of smaller agglomerated spherical NPs ofapproximately 40 – 50 nm diameter. EDX analysis of thisagglomerate showed that each region consisted of Ag and S
with varying ratios. Two regions in the 100 – 200 nm
aggregate had a Ag/S ratio of 2:1 (spot 1 and 2), whereas

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Figure 1 (See legend on next page.)

Page 4 of 18

Doolette et al. Chemistry Central Journal 2013, 7:46
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Page 5 of 18(See figure on previous page.)
Figure 1 Silver concentrations in the effluent and the total mass of Ag added to the AgNP dosed (a); Ag+ dosed (b); and, control(c) SBRs. Less than 1% of added Ag was found in the effluent. Nominal Ag concentrations were calculated from measured Ag spiking solution
concentrations. Error bars represent one standard deviation (n = 3).

one region contained Ag/S with a 1:1 ratio (spot 3)(Figure 3). The specific cause of NP aggregation requires
further investigation. However, it may be attributable to anumber of factors including the ionic strength of the mixed
liquor and the presence of organic chlorides and minerals.
Ionic strength in domestic wastewater is typically < 0.1 M,
whereas in anaerobic digesters IS is < 1 M, this may
cause NP homocoagulation (see Additional file 1: Table
SI.2 for wastewater elemental analysis). Conversely
heterocoagulation of NPs may arise from the interactionof AgNPs with organic chlorides and minerals.
Silver sulfide (as α-Ag2S) in the nano- size range has
previously been identified in sludge [8,12]; however, todetermine the crystal phase of the nano-sized particles
in this study, further crystallographic investigation is required. At temperatures < 173°C, the monoclinic crystalline
form of silver sulfide (α-Ag2S) dominates (acanthite). For
this phase to exist in the current study, Ag(0) in the original
AgNP must be first oxidised to Ag+ [25,26].
A recent study [25] provided evidence for the directconversion of AgNPs to Ag2S via an oxysulfidation
mechanism which was dependent on the presence ofsmall amounts of dissolved O2. In the SBR experiment,
residual O2 did remain during the 110 min anoxic phase.
Interestingly, EDX analysis showed that S was present in
all nanosized Ag particles identified in the sludge despite
the very short anoxic phase relative to the sulfidation
reaction times of AgNPs (i.e. > 5 h [8]). Given the very
short anoxic phase (105 min), the results suggest thata)sulfidation of AgNPs in mixed liquor may occur more
rapidly than previously shown. Alternatively, AgNP
sulfidation in the SBR may have been a gradual processthat occurred during successive anoxic phases; i.e.
during each anoxic phase a fraction of AgNPs may
have been sulfidised until all AgNPs were sulfidised.
This pathway is unlikely though as ‘fresh’ AgNPs wereadded each day.
To the best of our knowledge, the results are the firstto identify sulfidised Ag nanosized aggregates in aerobically generated sludge. A previously study that identified
Ag2S in ‘aerobic’ mixed liquor of a pilot WWTP was nottruly representative of an aerated sample because themixed liquor was first subjected to anaerobic treatment[8]. We therefore suggest that Ag2S identified in thataerobic sludge would have been produced during theinitial anaerobic treatment because once formed, Ag2S
is very resistant to oxidation and dissolution of Ag(analogous to other metal oxides [27]). Overall, theresults show that in WWTPs, the sulfidation of AgNPs
may occur during activated sludge treatment prior toanaerobic digestion.
Silver speciation in wastewater, activated sludge andanaerobic digestate as determined by synchrotron
studies

Principal component analysis (PCA) and target transformation identified six standard compounds suitable for thefitting of Ag in sludge samples: Ag2S NPs, Ag-acetate,

b)
2
3

1

map

1 µm

Figure 2 Characterisation of transformed Ag nanoparticles in aerobic sludge samples using STEM-HAADF. (a) STEM-HAADF image of atypical NP aggregate containing sulfidised nanoscale Ag particles. The bright aggregates are indicative of high Z elements (b) STEM-HAADF
image of a typical Ag aggregate, particles 1–3 were characterised using energy dispersive X-ray (EDX) analysis (see Figure 3). The sludge sample
was collected from the AgNP spiked SBR at the end of the experiment SBR experiment.

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Page 6 of 18

1.0

1.0

Aerobic

1.0

Aerobic

Aerobic

Spot 1
Spot 2

0.8

Spot 3

C

0.8

Ag

0.8
Ag

Cu

C

0.4

0.2

Cu

Maximum Spot Intensity

0.6

Cu

Maximum Spot Intensity

Maximum Spot Intensity

S

0.6

S
Ag
0.4

0.2

0

2

4

6

X-ray Energy (keV)

8

10

S

C
0.4

0.2

0.0

0.0

0.6

0.0

0

2

4

6

X-ray Energy (keV)

8

10

0

2

4

6

8

10

X-ray Energy (keV)

Figure 3 Energy dispersive X-ray (EDX) spectra of sulfidised silver nanoparticles in aerobic sludge. Spectra were collected from thespecific spots indicated in Figure 2b.

Ag-glutathione (Ag-GSH), Ag-thiosulfate, Ag-foil (Ag0)and Ag2S (Figure 4). Examination of the XANES spectraof the six target compounds (Figure 4) showed that
Ag2S NP and Ag-GSH were very similar (also see XANES
difference spectra Additional file 1: Figure SI.2b). Therefore,
for these Ag standards, their percentage contributions tothe sample model fits were combined (Table 1). The tworemaining Ag-S models (Ag-GSH and Ag-thiosulfate) arenot easily distinguishable from each other by visualinspection of the spectra; however, examination of thedifference spectra does show considerable variation(Additional file 1: Figure SI.2b). Furthermore, the identities
of Ag-GSH and Ag-glutathione (Ag-GSH) standards wereconfirmed by EXAFS analysis (data not shown).
The Ag K-edge XANES spectra of all sludge samples
and samples from the wastewater experiment are shownin Figure 4 together with the six references that wereused in the linear combination fitting (LCF). The LCF
analysis provided good fits to all experimental data
(Additional file 1: Figure SI.5). Results show that thedominant Ag species identified in all aerobic and anaerobic
sludges was Ag bonded with sulfur (S). The contributionsof each standard varied between the different Ag treatments (Ag+ vs. AgNP) and also between each treatment
process (aerobic vs anaerobic) (Table 1). The exception tothis was in the wastewater samples.
The two spectra of AgNP dosed sludge shows that
AgNPs were completely transformed during the SBR
experiment and again during anaerobic digestion (Figure 4,
Table 1). To the best of our knowledge this is the first timethat sulfidation of AgNPs has been reported in aerated
sludge. The aerobic sample (labelled ANP) was dominated(85%) by sulfidised Ag species with minor amounts ofelemental Ag (15%). However, in the anaerobic sample(NNP), elemental Ag was not a significant component.
The absence of Ag(0) in the anaerobic AgNP sample issupported by analysis of the corresponding EXAFS spectra(Additional file 1: Figure SI.7) where Ag – Ag bonding was
not detected (Table 2).
In the anaerobic AgNP treated sludge, bulk Ag2S wasdetected (13%) whereas in the aerobic sludge it was notdetected by XAS. This suggests that the anaerobic digestion
process in WWTPs may be vital for the conversionof nano-sized aggregates to bulk forms. The spectral
differences that were observed between bulk Ag2S and
Ag2S NPs are a common feature when comparing the
XANES spectra of NPs and the bulk. Such effects havebeen reported for numerous NPs, including Au, CdS,ZnO and Fe2O3 NPs [28-30]).
Similar sulfidation trends are apparent for the Ag+
dosed sludge; the major components of the aerobic
and anaerobic sludges were Ag-S coordinated species(80% and 86%, respectively).This was supported by
EXAFS analysis which identified Ag– S bonding ineach sludge sample (Table 2). The major solid phase
speciation changes between aerobic and anaerobic
AgNP dosed sludge was the decrease in Ag(0) (15% to 0%)
and Ag2S NP (40% to 14%) and the subsequent increase inbulk Ag2S (8% to 30%). It is most likely that metallic Ag
was produced in the Ag+ dosed aerobic sludge by photocatalytic reduction of AgNO3, or alternatively by reducing
agents in the mixed liquor (e.g. hydrogen sulfide, glucose).
Comparison of the XANES spectra for anaerobic Ag+
spiked sludge and anaerobic AgNP sludge shows only
minor differences. However, much greater differenceswere found between the aerobic AgNP treated sludge andthe anaerobic Ag+ sludge (Additional file 1: Figure SI.6).
Silver acetate was identified as a significant component(as determined by the size of the residual following leastsquares refinement of the model compounds during LCF)
in the XANES fitting of the anaerobic AgNP sludge (8%)
but not in the aerobic sample. The structure of this

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Page 7 of 18

standard was not verified by other methods, however, thespectra is significantly different from the Ag-S and Ag(0)standards (Figure 4 & Additional file 1: Figure SI.2b) to
be confident that Ag-carboxyl groups are present in theanaerobic sludge sample. In samples where Ag-acetate was
detected as a significant component (Table 1), re-fitting thespectra with Ag2O produced a poorer quality fit with larger
R2 values. Furthermore, when Ag-acetate was excluded
from the model, this led to an increase in the fit residuals(see Additional file 1: Table SI. 3 for increased residual
values).
Overall the majority of Ag in the AgNP dosed aerobic
and anaerobic sludges was sulfidised (85% and 92%,
respectively). The results are in agreement with previous
studies that have shown sulfidation of AgNPs in sludge[8,12] and highlights the importance of considering Ag
speciation in determining the fate and toxicity of AgNPs
in terrestrial environments.
The speciation of AgNPs in influent wastewater has
not been previously investigated. Based on Ag K-edge
XANES results, the results demonstrate that the absenceof activated sludge in influent wastewater had a considerable effect on AgNP transformation (Table 1). There
was complete transformation of AgNPs to Ag(0) for all
wastewater samples; no other species was identified as asignificant component in the fit model. Furthermore,
there were only subtle differences in the spectra forsamples collected initially (4 min after spiking) and after
3.5 h. (Additional file 1: Figure SI.4). The results suggestthat when PVP coated AgNPs enter wastewater, theirpolymer coating will be quickly modified or lost, andaggregation will occur. Additional analysis using the
PVP AgNP reference in place of metallic Ag(0) as atarget component, produced a poorer fit with a greater
residual (3.14 cf. 0.72). It should be noted that the PVP
coating of the AgNPs used in the experiments has notbeen fully characterised (i.e. coating thickness), so thiseffect may not be observed for all PVP coated NPs.
However, it can be concluded that that when AgNPs
enter WWTPs, the polymer surface coating may
already be modified and AgNPs will no longer be nanoin size. The size increase may be caused by a numberof factors, including heterocoagulation with natural
colloids (e.g. dissolved organic matter and iron andmanganese oxyhydroxides) and aggregation due to highionic strength of the wastewater.
Figure 4 Silver XANES K-Edge spectra of sludges and the 6 Ag
references used for the PCA. Where ANP = aerobic sludge fromthe AgNP dosed SBR; NNP = anaerobic sludge from the AgNP
treatment; NI = anaerobic sludge dosed with Ag+; AI = aerobic
sludge from the SBR dosed with Ag+; Ag-thio = Ag thiosulfate
complex; and, Ag-GSH = Ag glutathione complex. For spectra of thecontrol sludge (collected from the SBR that was not spiked with Ag)
see Figure SI.4.

Effect of silver nanoparticles on nitrification

The ammonium (NH+4 ) and nitrate/nitrite (NOx) profilesof each SBR are illustrated in Figures 5 & 6. There wasnear complete removal of NH+4 (> 99%) observed fromeach SBR during each cycle. It should be noted that forthe cycle analysed on day 2 for the AgNP dosed SBR only
70% of NH+4 was removed. In a similar SBR experiment,

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Page 8 of 18

Table 1 Linear combination fitting analysis of XANES spectra of sludges collected from the SBRs (aerobic), sludges
after anaerobic batch tests (anaerobic) and from the short term wastewater experiment
SampleAg2S

Ag2S NP + Ag-GSH

Ag(0)Ag-acetate

Ag-thiosulfate

Residual

40 (5)

20 (3)

19 (3)

23 (4)

0.350

72 (8)

6 (0.9)14 (0.9)85 (4)

15 (4)

Aerobic sludges
Control
+

Ag

8 (2)

AgNP

0.029
0.022

Anaerobic sludges
Control

39 (9)

24 (7)

27 (4)+

11 (6)

0.549

Ag

30 (1)

56 (2)

13 (0.7)

0.015

AgNP

13 (2)

78 (6)

8 (0.8)0.030

Wastewater Experiment
4 min

100 (0.2)

0.776

24 min100 (0.2)

0.863

210 min

100 (0.2)

0.927

The proportion of species are presented as percentages with the estimated standard deviation (SD) in parentheses. For the Ag2S NP + Ag-GSH column, the SD is
the sum of the individual SD’s from each species. The control sludge was collected from the SBR that was not spiked with Ag.

slight inhibition of nitrifying organisms by AgNPs(citrate capped AgNPs, 0.1 mg Ag L-1 of mixed liquor) was
also observed on the first day of Ag addition [9]. However,the current results are most likely due to unexpected
incomplete mixing of mixed liquor. In the cycle immediately following, complete mixing was resumed, and thus
complete NH+4 removal would be expected as occurred in
the other SBRs for day 2. At the beginning of the aeration
phase the highest concentrations of NH+4 were observed,
Table 2 Structural parameters of sludges and standards
derived from EXAFS analysis
Sample

Shell

CNa

Rb

σ2c

Ag – S

2.0

2.55

0.007

Aerobic sludges
Ag+

Ag – S

1.6

2.51

0.005

Ag – Ag

2.4

2.89

0.005

Ag – Ag

3.0

3.09

0.006

Ag – Ag

2.4

4.98

0.003

Ag+

Ag – S

2.0

2.48

0.008

AgNP

Ag – S

2.0

2.51

0.005

AgNP

Anaerobic sludges

References
Ag2S NPAg – S

1.5

2.52

0.004

Ag – Ag

3.0

3.06

0.007

Ag-foil

Ag – Ag

12.0

2.86

0.001

Ag-GSHAg – S

2.0

2.49

0.003

Ag – Ag

1.0

3.03

0.007

Ag-thiosulfate

Ag – Ag

1.0

2.52

0.002

a

Coordination number, b Interatomic bond distance (Å), c Debye-Waller factor.

with low variation between each SBR. The maximum concentrations were recorded on different days for the control(day 4; 24.5 mg NH+4 L-1), AgNP (day 6; 20.8 mg NH+4 L-1)
and Ag+ (day 3; 20.3 mg NH+4 L-1) dosed SBRs. Small
amounts of nitrite (NO-2) were produced in the reactors,
however, even during the cycles that had the highest
concentrations, NO-2 was completely converted to nitrate(NO-3) before the end of the phase.
Nitrification rates were calculated using linear regression over time for two complete cycles on days 8 and 9
and normalised for TSS content (Table 3). The rates
were calculated from the initial reduction of NH+4 at the
beginning of the aeration phase (Figure 5). Nitrification
occurred rapidly in the first 50 – 60 min of the aerobic
phase, and as a result the linear regression is based on 3
time measurements. To support these results, an on-line
NHx autoanalyser (YSI, USA) was also used on days 8
and 9 to measure NH+4 concentrations in the AgNP and
Ag+ dosed SBRs, respectively (Additional file 1: Figure SI.1).
A comparison of the nitrification rates calculated from bothanalysis methods shows comparable results (Table 3);
confirming the accuracy of the chemical data. Probe
determined NH+4 concentrations are between 4.5 mg L-1
and 5.4 mg L-1 lower than those obtained from chemical
analysis (Table 3). This may be due to the close proximityof the probe to the aeration stone in the SBR. Alternatively,
the lower pH of the mixed liquor compared to thecalibration solutions may have caused a shift of the NH+4
equilibrium (NH3 + H+ ⇌ NH+4 ) to the left, decreasing theconcentration of NH+4 . There was very limited variation in
the nitrification rates of all SBRs.
Several studies have investigated the effect of AgNPs
on nitrification in WWTPs, but results are conflicting

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40

40

40

Page 9 of 18

SBR 1- Control[NH - N]

[NO3 - N]

30[NO2 - N]
Aeration phase

20

10

0

Time + date

SBR 2 - Ag NP[NH4 - N][NO3 - N]

30
Aeration phase[NO2 - N]

20

10

0

Time + date

SBR 3 - Ag+[NO3 - N][NH4 - N]

30
Aeration phase[NO2 - N]

20

10

0

Figure 5 NH+4 and NOx profiles of each SBR. Results are shown from day 2 to day 9 (after spiking). Sample collection commenced on day 1;
the second day of SBR operation (results not shown). Samples were collected daily during one cycle (4 cycles in 24 h).

Time + date

Doolette et al. Chemistry Central Journal 2013, 7:46
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Page 10 of 18

a)

b)

c)

Figure 6 Variations of (a) NH+4 - N, (b) NO3 – N and (c) NO2 - N profiles during one complete 6 h cycle on day 9. Where D and F are thedecant and feed phases, respectively.

[9,14,15,18]. As the sludge matrix is likely to have amajor influence on the fate of AgNPs in WWTP, the results from the current experiment are most comparable
to those experiments that have used WWTP sludge andactivated sludge [9,15]; not synthetic wastewater. In a 15
d simulated SBR experiment [9], NH+4 removal efficiency
was not affected by AgNPs in wastewater (0.5 mg Ag
L-1), whereas in a short term (12 h) batch test using a
synthetic feed solution [15], a 7% decrease in nitrification rate at 1 mg AgNP L-1 was recorded. This inhibition
may be due to the relatively high DO concentrations(~ 7.2 – mg L-1) compared to the more realistic concentrations used in our experiment (1.5 – 2.5 mg L-1).
In the current experiment, the complex sludge matrix
may have decreased AgNP toxicity for a number ofTable 3 Nitrification rate for each SBR on days 8 and 9 ofthe experiment
Day 8

Day 9

SBR

mg L-1.h

mg L-1.hControl

2.4

1.8

+

Ag

1.5

1.3 (1.2)AgNP

2.6 (2.2)

1.6

Rates calculated from the on-line NH4 probe are shown in parentheses. All
rates have been normalised for TSS content.

reasons, with two primary factors being the presence of organic matter and the high ionic strength. Organic mattercomplexes Ag+ [31,32], which has been linked to AgNP
toxicity, whereas high salt concentrations cause NP aggregation which is known to decrease nanoparticle toxicity[33]. In addition, Ag+ anion binding may produce very
stable products such as AgCl (Ksp, H2O, 25°C = 1.77 x 10-10)
and Ag2S (Ksp, H2O, 25°C = 5.92 x 10-51) which will also
decrease Ag+ bioavailability and hence toxicity.
Effect of silver nanoparticles on methane production

The cumulative production of biogas (methane) during
anaerobic digestion of the AgNP and Ag+ dosed sludges
is shown in Figure 7. Based on the calculated anaerobic
biodegradability parameters, AgNPs did not have an
impact on sludge digestion (Table 4). There was nodifference between the methane production of AgNP, Ag+and control sludges at Ag concentrations of 184, 77 and6.3 mg Ag kg -1. The results concur with previous studies
that found methanogenesis was not affected by AgNPs atconcentrations < 18.9 mg Ag L-1 [18] and 40 mg Ag L-1[19]. Similarly, for bulk Ag, the rate and extent ofmethanogenesis in mixed cultures was not affected byeither AgNO3 or Ag2S at concentrations of 100 mg Ag L-1[34]. To the best of our knowledge, the results are the-1

Cumulative methane production (ml CH4 gVS )Doolette et al. Chemistry Central Journal 2013, 7:46
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Page 11 of 18

300

250

200

150

100

AgNP
Ag+
Control

50

0
0

10

20

30

40

Time (day)Figure 7 Cumulative methane produced during the 38 d anaerobic digestion. All values are blank corrected and the error bars show 95%
confidence intervals calculated from triplicate measurements.

first to demonstrate that transformed AgNPs in sludges(present mainly as Ag bonded to S groups) (Table 1), as
opposed to ‘pure’ AgNPs, have no effect (at 184 mg Ag kg-1)
on methanogenic processes which are essential forsludge degradability in WWTPs.
Effect of silver nanoparticles on niche microbialcommunities

Following mixed liquor digestion in the three SBRs andanaerobic assays, the diversity of bacterial populations wasdetermined and compared to that of influent wastewater,
activity sludge mixed liquor (aerobic inoculum) and anaerobic inoculum. The results from a PCA of the individual
data sets shows that all samples could be grouped based on
their source (Figure 8). In each case, there was a slight shift
from control/Ag+ to AgNP.
One dimension could account for 70% of overallvariation. Aerobic samples were heavily dominated by amajor dominant γ-Proteobacteria 19%, 22% and 21% for
the control, AgNP and Ag+ aerobic samples, respectively(Figure 8) This is surprising as nitrifiers and phosphate
accumulating organisms (PAOs) in activated sludge areusually β-Proteobacteria with only a small percentage
Table 4 Anaerobic biodegradability of each sludge as
indicated by degradation extent (fd,), apparent first orderhydrolysis rate coefficient (khyd) and the estimatedmethane potential (B0)Treatment

khyd (d-1)

fd

B0 (mL/gVS)

Control

0.13 ± 0.020

0.31 ± 0.016

195 ± 1Ag+

0.12 ± 0.014

0.36 ± 0.014

228 ± 9AgNP

0.12 ± 0.014

0.36 ± 0.014

238 ± 9from the gamma subclass. Removal of organic N is a
two-step process where ammonia is initially oxidised tonitrite by ammonia oxidising bacteria (AOB) and thenfurther oxidised to nitrate by nitrite oxidising bacteria(NOB). All AOB belong to two genera each in the
β-Proteobacteria and γ-Proteobacteria phylum, whereas
NOB belong to five genera in various classes of theProteobacteria. The primary habitats of γ-Proteobacteria
are marine environments whereas β-Proteobacteria
dominate in freshwater systems. The inoculum plant
was in a coastal environment (more saline), which could
account for this increased dominance.
Ammonia oxidising bacteria are generally more sensitive
to toxicants than NOB [16,35]. There was a very minor
response to the AgNP treatment, driven mainly by slightshifts in niche populations. Subdominant microbial
structure in the Ag+ treated sludges was not significantly different to that of the control. What was more
surprising was that there was almost no shift in population between the inoculum and Ag+ and control. This is
surprising as the feed is different, the mode of operation isdifferent (continuous in parent vs sequenced in SBR),
and at least 1 nominal sludge age occurred throughthe study. The pyrosequencing data confirms the nitrification results, in that AgNPs at a concentration of
2.5 mg Ag L-1 (2.9 g TS kg-1), do not influence thebroad microbial population.
Anaerobic samples indicated that control and Ag+ werevery similar, but with a large shift from inoculum to batch,and a small shift from control/Ag+ to AgNP (Figure 9).
This was confirmed through additional PCA analysis on
the anaerobic samples only (top 500 OTUs, Hellinger adjusted). This indicated a large shift from inoculum to endDoolette et al. Chemistry Central Journal 2013, 7:46
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Page 12 of 18

0.5

FEED

Streptococcus

0.4
0.3
Clostridiales

0.1

AD inoc
Methanosaeta
OP8
Spirochaetes
Methanobacteriales

AD
0

(16%)

PC2 (16%)

0.2

ASActinomycetales
Saprospiraceae

-0.1

WS6

-0.2
Gammaproteobacteria
-0.3
-0.4
-0.5
-0.5

-0.4

-0.3

-0.2

-0.1

0
0.1
PC1 (52%)

0.2

0.3

0.4

0.5

Figure 8 Overall bi-plot of PCA data showing top 10 OTUs. All OTUs were used for analysis. The key clusters of Activated sludge - AS
(inoculum Ag+, and control overlap), Feed, and Anaerobic Digestion (AD) are shown. Arrows show shift from control and Ag+ to AgNP
communities. Note that AD Inoculum is right-shifted on PC1 compared to the control and Ag+ samples.

BMP, with a dominant WS6 OTU being largely replacedby OP8 (both candidate divisions), and a number of othermajor OTUs. Silver NPs seemed to cause a subtle shift
from Spirochaetes to other organisms. Archaea seemed
not at all influenced by batch operation, or Ag+/AgNP
treatment. Our data therefore support those in previous
studies [19] indicating no impact of AgNPs on Archaea
compared to controls (at 20 mg AgNP L-1).

What is surprising is the dominance in all anaerobic
samples by uncultured division OTUs (Figure 9). The
inoculum contained phyla from the uncultured candidate
division WS6 (18%), whereas microbial population inanaerobic samples collected after digestion appeared to
be dominated (10 – 17%) by organisms from anothercandidate division (OP8). The cause of this shift in diversity
is unclear. The WS6 phyla was first identified in a
Spirochaetes

0.4

0.3

0.2
WS6' Spirochaetes
Control
Bacteroidales
Methanobacteriales

PC2 (10%)0.1

Gammaproteobacteria
WS60

Lysinibacillus
Spirochaetes

Saprospiraceae
Saprospiraceae

Inoc.

Ag+
OP11'
Methyloversatilis
Cytophaga

Clostridiales
Thermovirgaceae OP8
Verrucomicrobiales
Dethiosulfovibrionaceae'
Ag Nan

-0.1

-0.2

-0.3

-0.4
-0.5

-0.4

-0.3

-0.2

Figure 9 PCA analysis of anaerobic pyrotag results only.

-0.1

0PC1 (80%)

0.1

0.2

0.3

0.4

Doolette et al. Chemistry Central Journal 2013, 7:46
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contaminated aquifer and has since been identified inother environments (e.g. anoxic pond sediment [36],sulfur-rich spring sediments [37], eutrophic estuaries[38], hydrothermal vents [39]) but not in sludges orwastewater. Organisms from the OP8 division havebeen identified in mangrove sediments [40] and ananaerobic sludge digester, where 1% of the operationaltaxonomic units (OTUs) were represented by organisms
from this division [41]. Broadly speaking, environmental
bacterial community structure is regulated by local
conditions. Hence, in the BMP test, factors including
salinity and nutrient conditions [42] may have differedto that in the tank where inoculum was collected causing
the population change.
Sensitive methanogenic microbes (Methanosaetaceae)
accounted for ~ 11% of the variation in all anaerobic
samples, including the inoculum, and was not impacted
by the presence of AgNPs (Figure 9). This organism ismost sensitive to possible surface active agents [21]. It ishighly important that the bacterial population changed
so strongly between inoculum and the end of the batch,while the archaeal population seemed untouched. Thismeans that the mode of operation has a strong impact
on acidogenic microbial populations but not methanogenic ones. It will be important to further evaluate therole of organisms in candidate divisions, as almost nothing is known of these microbes.
Whilst previous studies have investigated the effects ofpure AgNPs on wastewater microbial populations, this is
the first time that the influence of transformed AgNPs(primarily as Ag-S species) has been investigated.
Moreover, to the best of our knowledge, this is also thefirst time that transformed AgNPs have been shown toaffect niche populations. The results cannot conclusively
be attributed to a nano- effect due to the higher spike
concentration of Ag in the AgNP treated sludge
compared to the Ag+ sludge. Yet, XAS analysis of theanaerobic sludges did show a greater percentage of nanosized Ag2S in the AgNP treated sludge compared to Ag+
dosed sludge (78% cf. 53%, respectively) and a lower
percentage of bulk Ag2S (13% cf. 30%, respectively). This
supports our hypothesis that the observed population
changes are attributable to a nano-effect, althoughfurther research is required to confirm this hypothesis.
Nevertheless, the results still demonstrate that even after
their transformation to much less toxic Ag-S species,
AgNPs have the potential to impact niche microbialcommunities but are not likely to impact overall WWTP
microbial processes (e.g. nitrification and methanogenesis).

Conclusions
In our experiments, > 99% of PVP-coated AgNPs were
removed from wastewater when subjected to activatedsludge digestion. During the SBR experiment andPage 13 of 18

subsequent anaerobic digestion stage, nitrogen removal
and methane production (respectively) were not affectedby transformed AgNPs.
Pyrosequencing analysis of microbial communities
showed that AgNPs and Ag+ did not affect the dominant
populations of nitrifiers and methanogenic organisms inaerobic and anaerobic generated sludges, respectively.
However, in both sludges a subtle shift in niche populations
was observed. In the case of aerobic sludge samples, theshift was extremely minor, whilst for anaerobically digested
samples there was a much larger shift. Additional studies
are required to confirm if this change in population is
exclusively a nano- effect.
Two conclusions were drawn from the XAS analysis ofsludge: (i) AgNPs were sulfidised during SBR operation
followed by near complete sulfidation during anaerobic
digestion; and (ii) AgNP dosed anaerobic sludge contained
a higher fraction on nano sized Ag2S species compared toAg+ dosed sludge. The production of stable Ag-S species
may have limited the toxicity of AgNPs towards nitrifiers
and methanogenic bacteria as Ag+ is believed to be themain toxicity mechanism of AgNPs.
Based on our results, PVP-coated AgNPs will notaffect nitrification and methanogenesis during WWT,
even in the future with the increasing use of AgNPs.
Further investigations are required to confirm whether
sub-dominant microbial sludge populations are at risk fromAgNP exposure, as this may have long term consequences
for the successful operation of WWTPs.

Methods
Preparation and characterisation of nanoparticle stock
solutions

Polyvinylpyrrolidone (PVP) coated (0.1%) Ag NP powders
were purchased from Nanostructured & Amorphous
Materials, Inc. (Houston, TX) (10 nm nominal particle
size diameter). PVP coated NPs were chosen as they are avery common AgNP capping agent. Thus, their use isrealistic of the AgNPs that would be released into wastewater streams. Stock suspensions of AgNPs were preparedby adding the NP powder (0.1 g) to ultrapure deionised
water (50 mL, 18.2 Ω) and sonicating (1800 W, 3 min) in
an ice bath. The AgNP suspension was then centrifuged(2200 g, 15 min) producing a final stock suspension
with an average Ag concentration of 35.7 mg Ag L-1(SD = 5.6 mg Ag L-1, n = 11), 8% of which was dissolved
ionic Ag+ [43]. The AgNP stock suspensions were prepareddaily (30 – 60 min before spiking). Total Ag concentrationsof the NP spiking solutions were determined by an
open-vessel acid digestion (HNO3, 70%) method as
described previously [43].
The AgNP suspensions prepared with this method
have been extensively characterized previously using
dynamic light scattering (DLS, Malvern Zetasizer),Doolette et al. Chemistry Central Journal 2013, 7:46
http://journal.chemistrycentral.com/content/7/1/46

transmission electron microscopy (TEM, Phillips
CM200 at 120 keV) and X-ray diffraction analysis [44].
The particle size distribution has also been investigatedusing disk centrifuge analysis (CPS Instruments disccentrifuge 24000 UHR). In summary, the average particle
diameter was between 40 nm with < 8% of Ag existing as
dissolved Ag+ (Additional file 1: Table SI.1 for complete
NP characterisation) [44]. Previous work [44] (using thesame method and batch of nanoparticles) has showed thatthe volumetric diameters of the AgNPs observed using
TEM corresponded with the crystallite size determinedfrom X-Ray diffraction analysis, the hydrodynamic
diameter (dh) observed using DLS and the Stokes diameter as found using disk centrifugation.

Set-up and operation of sequencing batch reactors

Three individual SBRs (control, Ag+ and AgNPs) with aworking volume of 5 L and an initial TSS of 6.0 g L-1
were prepared with 0.9 L of activity sludge mixed liquor(TS = 35.3 g L-1) and 4.1 L of influent wastewater. Activatedbiological nutrient removal (BNR) sludge was collectedfrom an activated sludge wastewater treatment plant(Luggage Point), and influent wastewater was collectedfrom a local domestic wastewater pumping station(Indooroopilly), both located in Brisbane, Queensland,
Australia.
Each SBR was operated with four 6 h cycles per daywith a hydraulic retention time (HRT) of 15 h. Eachcycle consisted of a 3 h aerated aerobic phase, followed
by 50 min settling, 15 min decanting, 10 min feeding
and 105 min anoxic [low dissolved oxygen (DO)] phases.
Feeding, decanting and sampling ports were located at
different positions on the reactors. During the 3 h
aerobic stage, DO levels were maintained between 1.5 –2.5 mg L-1 by intermittent aeration, controlled with an
online DO detector. Following the settling phase, 3 L ofsupernatant was decanted and 3 L of cold influent wastewater (20°C) was pumped into each SBR. The reactors
were continuously stirred with a magnetic stirrer except
during settling and decant phases.
Silver (as NPs or AgNO3) was added once every 24 h
at the beginning of a feed cycle and for the remaining
three feed cycles in that 24 hours no Ag was added.
Prior to spiking, trace amounts of Ag were recordedin the mixed liquor of each SBR (day 0), (36, 26 and24 μg Ag L-1 for the control, Ag NP and Ag+ dosed
SBRs, respectively).
Following the 10 day aerobic digestion, sludge wasallowed to settle for 2 h and the supernatant decanted.
The remaining sludge in each SBR was centrifuged(2 min, 3250 g), to increase the TS concentrations(Table 5), and approximately half was used in the subsequent anaerobic digestion experiment.

Page 14 of 18

Table 5 Average characteristics of each sequence batch
reactorCharacteristic

Control SBR

Ag NP SBR

Ag+SBR

Average pH

6.71 (0.23)

6.67 (0.07)

6.62 (0.09)

4.4

4.5

6.3

Final TSS (g L-1)
-1

Final VSS (g L )

4.0

4.3

5.5

TS after centrifugation of SBR
sludge (g kg-1)

72.8

60.5

65.4

Sludge [Ag] before spiking(mg Ag L-1)

0.04

0.03

0.02

Total Ag added (mg)

0.00

12.73

6.14

Mixed liquor suspended and volatile solids (MLSS and
MLVSS, respectively) were analysed every 2 d according
to APHA Standard Methods (1992). The chemical
characteristics and Ag spiking concentrations of each
SBR are given in Table 5.
Transmission electron microscopy analysis of silver
nanoparticles in sludge

Freeze dried sludge was collected at the conclusion of
the 10 d SBR process to determine whether physical orchemical transformation of AgNPs had occurred in the
AgNP dosed SBR. STEM analysis in HAADF mode wasused to investigate the morphology of AgNPs in thesludge, whereas EDX together with TEM was used forelemental analysis. Samples of aerobic sludge werecollected at the end of the SBR experiment. Samples
were freeze-dried and analysed using a JEOL 2100 TEM
operating at 200 kV. Freeze dried samples were groundin methanol using a mortar and pestle and a single drop
pipetted onto a 200-mesh copper (Cu) TEM grid withlacey carbon support film and allowed to evaporate atroom temperature.
The elemental composition of “bright” NPs/aggregates
was investigated using an EDX spectrometer. The TEM
was used in scanning mode (STEM) with a high-angle
annular dark-field (HAADF) detector. The high angle
detector collects transmitted electrons that are scattered(primarily incoherently) to high angles, whilst excluding
Bragg (coherent) scattering. The detector provides animage where the contrast is dependent on the approximate
square of the atomic number (Z). Accordingly, brightspots in the image correspond to high Z elements;
making the detection of Ag containing particles in thecomplex sludge matrix more straight forward than thatin a bright-field image.
Solid phase speciation of silver in sludge usingsynchrotron radiation

Solid phase speciation of Ag in aerobic and anaerobic
sludges was further examined using X-ray absorption

Doolette et al. Chemistry Central Journal 2013, 7:46
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spectroscopy (XAS); specifically X-ray absorption nearedge structure (XANES) and extended X-ray absorption
fine structure (EXAFS) analysis.
Silver K-edge X-ray absorption spectra were recordedon the XAS beamline at the Australian Synchrotron(AS), Melbourne, Australia. The 3 GeV electron beam
was maintained at a current of 200 mA in top-up
during the sample analysis. The X-ray beam wastuned with a Si (311) monochromator in the energy
ranges of 25312–25492 eV for pre-edge (10 eV steps),
25492–25562 eV XANES region (0.5 eV steps) andthen 0.035 Å–1 steps in k-space for EXAFS. A metallic Ag foil, recorded in transmission mode downstream of the sample, was used as an internal
standard to calibrate the energy scale to the first peakof the first derivative of the Ag edge (25515 eV).
Spectra of the samples were recorded in fluorescencemode on a 100-pixel Ge detector array at 90o to theincident beam (Canberra/UniSys).
Freeze-dried sludges (aerobic and anaerobic) were
finely ground to a homogenous powder and compressed
into pellets with a hand press. Samples that had a highAg concentration were diluted with cellulose material(Sigma-Aldrich). All samples were cooled to ~10 K in a
Cryo Industries (Manchester, New Hampshire, USA)cryostat. One to eight scans per sample were collected
for XANES spectra and 14 scans per sample were collected for EXAFS spectra. Reference materials measured
at the XAS beamline included PVP-coated AgNPs(Nanoamor), AgNO3, Ag2S, Ag2O, AgCl, Ag2CO3, andAg2SO4 (all purchased from Sigma Aldrich). Additional
standards were prepared the day of analysis and stored
in the dark until use; Ag2PO4, Ag-goethite, Ag-kaolinite,
Ag-humic acid complex, Ag-fulvic acid complex, Ag
thiosulfate (STS), Ag-acetate, Ag-glutathione (Ag-GSH)
and Ag2S NPs. (See Supporting Information for synthesis
and preparation of all Ag standards).
Solid phase speciation of silver nanoparticles inwastewater using synchrotron radiation

A short term exposure experiment (3 h) was undertaken
to examine the potential rapid transformation/reactionsof AgNPs in wastewater (in the absence of activatedsludge) using synchrotron based XAS. PVP coated
AgNPs were spiked into wastewater (500 mL) to a final
concentration of 5.4 mg Ag L-1. The wastewater was thesame as that which was used in the SBR study. The
AgNP-wastewater suspension was stirred continuously
for 210 min and the DO concentration was measured
with an online DO detector (7.4 mg O2 L-1 to 7.1 mg
O2 L-1). Approximately one mL of the suspension wascollected at t = 4, 10, 24, 45, 60, 94 and 210 min after
the addition of AgNPs. Each sample was collectedusing a two mL glass syringe and injected directly into a

Page 15 of 18

leucite cuvette, secured with Kapton tape, immediately
flash frozen in liquid N2 and stored in dry ice until
XAS analysis.
XAS data analysis

The chemical speciation of each sample was determinedby fitting a linear combination of model compounds toeach XANES spectrum (Additional file 1: Figure SI.3).
The number of components in the sample XANES
spectra was determined using principal component
analysis (PCA) of all sample spectra, followed by target
transformation to identify the most likely componentsof the spectra from a model compound library. The
number of components to fit were chosen from theeigenvalues from the PCA and visual inspection of theplot of eigenvectors.
Linear combinations of the six spectra were fitted toeach sample spectrum with the best fit to the experimental spectrum achieved by least squares refinement of
the model compounds to the experimental spectrum.
The best fits were improved by the removal of small
components (< 1%) and the adequacy judged by the sizeof the residual and visual inspection to ensure that allfeatures were accounted for.
Calibration, averaging and background subtraction of
all spectra and principle component, target and multiple
linear regression analyses of XANES spectra were
performed using EXAFSPAK software package (G.N.
George, SSRL).
Investigation of nitrification inhibition and silver
partioning during aerobic digestion

For nitrification analysis, homogenous mixed liquor
samples were collected daily at the end of feed, anoxic, aerobic and settling phases during one 6 h cycle
for the first 7 d of SBR operation. On days 9 and 10,
more frequent sampling was conducted during the
aerobic and anaerobic phases. Samples were filtered(< 0.22 μm) and stored at 4°C until analysis.
The choice of Ag spiking rate was a compromise
between realistic environmental exposure concentrations[11], previous partitioning studies of Ag NP in wastewater[8,18] and instrumental detection limits. Taking thesefactors into account, the aim was to produce sludges
with a final concentration of ~100 mg Ag kg-1 TS for
the Ag+ and AgNP treatments.
The SBRs were operated for 24 h before spiking toallow for equilibration of the mixed liquor. Reactors 2
and 3 received the AgNP and Ag+ (as silver nitrate(AgNO3)) treatments, respectively. Reactor 1 was
assigned the control SBR and received ultrapure
deionised water (Millipore) at each spiking event inorder to maintain a consistent volume for all threereactors. Treatments were added once daily, for 8 d,

Doolette et al. Chemistry Central Journal 2013, 7:46
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to each SBR at the beginning of the aeration phaseby pipetting the spiking solution directly into thereactor. Each reactor received a total of 330 mL of
the assigned spiking solution (nominal concentration
for AgNO3 and AgNP suspensions = 20 mg Ag L-1).
The Ag concentration of the ultrapure deionised
water used in the control SBR was below the limit ofdetection for ICP-MS analysis (< 0.05 μg L-1). Reactors
were operated for a total of 10 d.
For silver analysis, mixed liquor (10 mL) and effluent(35 mL) samples were collected once daily from each
SBR; 3 and 5.5 h after spiking, respectively. Samples
were acidified and stored at 4°C before subsequent
digestion and analysis for total Ag by ICP-MS.
Anaerobic digestion and biomethane potential test

The effect of AgNPs on anaerobic digestion wasassessed using a biomethane potential (BMP) test as
previously described [45]. Anaerobic biomethane potential tests (BMP) were carried out for AgNPs andAg+ using sludge collected from each SBR at the conclusion of that experiment (all assays in triplicate).
The inoculum (activity sludge) had a broad trophic microbial composition to ensure the substrate would not belimited. Blank assays (in triplicate) were used to determine the background methane production from theinoculum.
Aerobically digested sludges from the SBRs (substrate)
were diluted to 30 g L-1 (TSS) with ultrapure deionised
water (Millipore). Substrate (40 g wet) and inoculum(60 g wet) (anaerobic digestate from a municipal WWTP,
Brisbane, Queensland) were added to glass serum bottles(160 mL working volume), flushed with high purity N2 gas
for 3 min (1 L min-1), sealed with a butyl rubber stopperand aluminium crimp-cap and stored in a temperature
controlled incubator (36°C) for 38 d. Blanks (n = 3)
contained inoculum (60 g) and ultrapure deionised water(40 mL) (Millipore). Each assay was performed in triplicate. Once daily methane production had ceased (38 d)
the batches were terminated, and analysis of the microbial community was conducted. Biogas volume wasmeasured periodically (initially daily) and the quality(CH4, CO2, H2) was analysed using gas chromatography, with a thermal conductivity detector (Perkin
Elmer). Confidence intervals (95%) were calculated
from triplicate measurements and were ≤ 0.02 g COD
d-1 for all samples. Excess CH4 was vented from each
serum bottle periodically via syringe and measured byliquid displacement.
Microbial community analysis: DNA extraction and 16 s
Pyrotag Analysis

To assess the potential impact of AgNPs on microbial
diversity of sludge samples (aerobic and anaerobic), a

Page 16 of 18

massive parallel sequencing approach using pyrotag
sequencing was used. Microbial diversity analysis was
conducted on samples of activated sludge mixed liquor(Luggage Point WWTP), feed (Indooroopilly pumping
station), sludge after aerobic digestion (control, Ag+ andAgNP), anaerobic inoculum (Luggage Point WWTP)
and anaerobic digestate (control, Ag+, and Ag NP).
Community genomic DNA from the anodic biofilms
were extracted using FastDNA SPIN for Soil kit (MPBiomedicals, USA) and Fastprep beadbeating machine(Bio101, USA) according to the manufacturer’s protocol.
The 3’ region of the 16S/18S rRNA gene was targetedusing universal primers 926 F (5’-AAACTYAAAKGAATTGACGG-3’) and 1392R (5’-ACGGGCGGTGTGTRC-3’).
Primer sequences were modified by the addition ofRoche 454 adaptor 1 or 2 sequences and unique 5 bp
barcodes at the 5’ end of the primer (sequences notshown) [46,47]. DNA concentration and purity wasthen determined by gel electrophoresis on 1% agarose
gel and spectrophotometrically using the NanoDrop
ND-1000 (Thermo Fisher Scientific, USA). DNA waslyophilised using Savant SpeedVac Concentrator SVC100H(Thermo Fisher Scientific, USA) and submitted to theAustralian Centre for Ecogenomics (ACE) for 16 s
rRNA gene pyrotag sequencing on the Genome
Sequencer FLX Titanium platform (Roche, USA).
Pyrotag sequences were processed using Pyrotagger[48], and QIIME with correction via ACACIA. Operational taxonomic unit (OTU) tables were normalised,and a square root (Hellinger) transformation was applied to emphasise comparison of niche populations
over dominants. A principal components analysis wasthen performed on the square root (Hellinger adjusted) normalised OTU table using Matlab (princomp
command), and results visualised using biplot.
Chemical analysis of silver spiking solutions, mixed liquor,
effluent, and sludge

Silver concentrations of the AgNP spiking solutions
were determined using an open-vessel acid digestion(HNO3, 70%) method as previously described [43].
Effluent and mixed liquor samples were analysed
for Ag following microwave digestion in aqua regia
according to the method used for wastewater previously [8]. Sludge samples (aerobic and anaerobic)
were first dried at 40°C to constant weight, and thenallowed to react with H2O2 before using the sameaqua regia digestion method. Silver concentrations inall digested solutions were determined using ICP-MS(Agilent 7500ce) and He(g) as a collision gas (4 mL min-1)
and monitoring Ag at m/z 107 and 109.
Filtered mixed liquor was analysed for NH+4 , NO-2
and NO-3 using a Lachat QuikChem8000 Flow Injection
Analyser.

Doolette et al. Chemistry Central Journal 2013, 7:46
http://journal.chemistrycentral.com/content/7/1/46

Additional file
Additional file 1: The following additional information data are
available with the online version of this paper in Additional file 1.
Methods for synthesis and preparation of synchrotron standards andsample preparation. Table SI.1. Characteristics of the silver nanoparticles(AgNPs) and AgNP stock suspensions. Figure SI.1. NH4 – N profiles of the
AgNP and Ag+ dosed SBRs as measured by the NH+4 on-line detector.
Table SI.2. The concentration of major and trace elements in the influent
wastewater Figure SI.2. Difference XANES spectra of sludge and various
Ag references used in LCF analysis. Figure SI.3. Ag K-Edge XANES spectraof all reference materials. Figure SI.4. Ag K-Edge XANES spectra of
aerobic and anaerobic control sludges and wastewater from the influent
experiment. Figure SI.5. Bulk silver (Ag) X-ray absorption near-edge
spectroscopy (XANES) of sludge collected from the SBRs (a-c) and after
the anaerobic batch test (d-f). Figure SI.6. Ag K-edge XANES spectra
showing the considerable difference between aerobic sludge dosed with
AgNP, and anaerobic sludge dosed with Ag+ or AgNP. Figure SI.7.
k3-weighted Ag K-edge EXAFS spectra of sludges and their respective
phase-corrected Fourier transforms. Table SI.3. The higher residual values
that resulted from the exclusion of Ag-acetate from the linear
combination fitting analysis of XANES spectra of sludges.
Abbreviations
AgNPs: Silver nanoparticles; TEM: Transmission electron microscopy;
EDX: Energy dispersive X-ray analysis; XAS: X-ray absorption spectroscopy;
TSS: Total suspended solids; TS: Total solids; MNM: Manufactured
nanomaterials; PVP: Polyvinylpyrrolidone; SBR: Sequencing batch reactor;
DOC: Dissolved organic carbon; WWTP: Wastewater treatment plant;
DLS: Dynamic light scattering; BNR: Biological nutrient removal;
DO: Dissolved oxygen; HRT: Hydraulic retention time; OTU: Operationaltaxonomic units; SD: Standard deviation; STEM: Scanning transmission
electron microscopy; HAADF: High angle annular dark field; PCA: Principal
component analysis; LCF: Linear combination fitting; XANES: X-ray absorption
near edge spectroscopy; EXAFS: Extended X-ray absorption fine structure;
AS: Australian synchrotron; Ag+: Dissolved ionic silver; Ag0: Elemental silver;
NH+4 : Ammonium; NO-3: Nitrate; NO-2: Nitrite; AgNO3: Silver nitrate; AgGSH: Silver glutathione complex; Ag2S NP: Silver sulfide nanoparticles; Agthio: Silver thiosulfate; Ag-HA: Silver – Humic acid; Ag-FA: Silver – Fulvic acid.
Competing interests
Authors declare that they have no competing interests.
Authors’ contributions
CD: designed and conducted the experiments, interpreted results and wrote
the manuscript. MJM: designed the experiments, interpreted results andparticipated in manuscript preparation. JK: designed the experiments,
interpreted results and participated in manuscript preparation. DJB: assisted
in experimental design, interpreted results and participated in manuscript
preparation. HHH: conducted XAS experiments, participated in XAS data
analysis, interpreted XAS results and participated in manuscript preparation.
HG: Set-up the SBR experiments, conducted the anaerobic batch test andassisted in analysis of results. GC: Participated in experimental design, wasinvolved in useful discussions and participated in manuscript preparation. All
authors read and approved the final manuscript.
Acknowledgements
We thank Kenn (Yang) Lu for his assistance in DNA extraction and quality
analysis, Fiona May at the Australian Centre for Ecogenomics (ACE) for
performing the pyrotag sequencing, Dr Beatrice Keller-Lehmann for FIA of
samples, Claire Wright (CSIRO) for ICP-MS analysis and Dr Graeme
Auchterlonie at the Centre for Microscopy & Microanalysis at The Universityof Queensland (CMM, UQ) for TEM/STEM/EDX analysis. This research wasundertaken on the XAS beamline at the Australian Synchrotron, Victoria,
Australia.
Author details
1
School of Agriculture Food & Wine, The University of Adelaide, PMB 1, Glen
Osmond, SA 5064, Australia. 2CSIRO Land and Water, Environmental
Biogeochemistry Program, Advanced Materials Transformational Capability

Page 17 of 18

Platform-Nanosafety, Waite Campus, Waite Road, Urrbrae, SA 5064, Australia.
Advanced Water Management Centre, The University of Queensland, St
Lucia, 4072, Queensland, Australia. 4School of Chemistry and Physics, The
University of Adelaide, Adelaide, SA 5005, Australia. 5Department ofChemistry, The University of Gothenburg, Kemivägen 10, 41296, Göteborg,
Sweden.
3

Received: 3 January 2013 Accepted: 28 February 2013
Published: 4 March 2013
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