Implications of Pony Lake Fulvic Acid for the Aggregation and

Subscriber access provided by Gothenburg University Library

Article

Implications of Pony Lake Fulvic Acid for the aggregation and dissolution of oppositely charged surface-coated silver nanoparticles and their ecotoxicological effects on Daphnia magna YounJung Jung, George Metreveli, Chang-Beom Park, Seungyun Baik, and Gabriele E. Schaumann Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.7b04635 • Publication Date (Web): 19 Dec 2017 Downloaded from http://pubs.acs.org on December 20, 2017

Just Accepted “Just Accepted” manuscripts have been peer-reviewed and accepted for publication. They are posted online prior to technical editing, formatting for publication and author proofing. The American Chemical Society provides “Just Accepted” as a free service to the research community to expedite the dissemination of scientific material as soon as possible after acceptance. “Just Accepted” manuscripts appear in full in PDF format accompanied by an HTML abstract. “Just Accepted” manuscripts have been fully peer reviewed, but should not be considered the official version of record. They are accessible to all readers and citable by the Digital Object Identifier (DOI®). “Just Accepted” is an optional service offered to authors. Therefore, the “Just Accepted” Web site may not include all articles that will be published in the journal. After a manuscript is technically edited and formatted, it will be removed from the “Just Accepted” Web site and published as an ASAP article. Note that technical editing may introduce minor changes to the manuscript text and/or graphics which could affect content, and all legal disclaimers and ethical guidelines that apply to the journal pertain. ACS cannot be held responsible for errors or consequences arising from the use of information contained in these “Just Accepted” manuscripts.

Environmental Science & Technology is published by the American Chemical Society. 1155 Sixteenth Street N.W., Washington, DC 20036 Published by American Chemical Society. Copyright © American Chemical Society. However, no copyright claim is made to original U.S. Government works, or works produced by employees of any Commonwealth realm Crown government in the course of their duties.

Page 1 of 32

Environmental Science & Technology

1

Implications of Pony Lake Fulvic Acid for the

2

aggregation and dissolution of oppositely charged

3

surface-coated silver nanoparticles and their

4

ecotoxicological effects on Daphnia magna

5

YounJung Jung†,‡, George Metreveli‡, Chang-Beom Park†, Seungyun Baik*,†, and Gabriele E.

6

Schaumann‡

7

†

8

Campus E 7.1, Saarbrucken 66123, Germany

9

‡

Environmental Safety Group, Korea Institute of Science and Technology (KIST) Europe,

Group of Environmental and Soil Chemistry, Institute for Environmental Sciences, University

10

of Koblenz-Landau, Fortstrasse 7, 76829 Landau, Germany

11

KEYWORDS: Silver nanoparticles (AgNPs), Pony Lake Fulvic Acid (PLFA), aggregation,

12

dissolution, ecotoxicological effects, Daphnia magna

13 14 15

ACS Paragon Plus Environment

1


16

Page 2 of 32

ABSTRACT:

17

Citrate (Cit) and polyethylenimine (BPEI)-coated silver nanoparticles (AgNPs) were used to

18

understand how the type of capping agents and surface charge affect their colloidal stability,

19

dissolution, and ecotoxicity in the absence/presence of Pony Lake Fulvic Acid (PLFA). In the

20

presence of PLFA, Cit-AgNPs were stabilized, while BPEI-AgNPs were aggregated. The

21

aggregation of BPEI-AgNPs decreased with the time and their stabilizing effect increased at high

22

PLFA concentration. The dissolution also differed between both AgNPs and was influenced by

23

the PLFA concentration. Generally, BPEI-AgNPs showed a lower amount of dissolved Ag than

24

Cit-AgNPs. The dissolved Ag concentration decreased for both AgNPs at low PLFA

25

concentration (5 mg/L). In contrast, the extent of nanoparticle dissolution increased at high

26

PLFA concentration (30 mg/L) but only for BPEI-AgNPs. In the absence of PLFA, the

27

ecotoxicity of Cit-AgNPs to Daphnia magna was higher than that of BPEI-AgNPs. However, the

28

ecotoxicity of AgNPs in the presence of PLFA was up to 70% lower than in their absence. We

29

demonstrated that the differences in colloidal stability, dissolution, and ecotoxicity may be

30

attributed to the different capping agents, surface charge and concentration of natural organic

31

matter (NOM) as well as to the formation of dissolved Ag complexes with NOM.

32 33 34 35 36


2

Page 3 of 32

37


INTRODUCTION

38

The increasing use of various products containing silver nanoparticles (AgNPs) has raised

39

concerns about discharging AgNPs into aquatic environments,1-4 due to their ecotoxicological

40

potential to aquatic organisms.5-8 Some researchers highlighted dissolved Ag released from

41

AgNPs as one of the main factors to induce adverse effects to the aquatic organisms.6,9 However,

42

other studies showed that AgNPs as well as dissolved Ag may play key roles in the ecotoxicity

43

of AgNPs.10-12 Therefore, it is worthwhile to study both the dissolution as well as the colloidal

44

state of AgNPs to better understand the relationships between their fate and ecotoxicity.

45

The fate of AgNPs in aquatic environments is determined by aggregation, sulfidation,

46

dissolution and adsorption onto various bio-geochemical interfaces and diverse results for

47

ecotoxicological impact are driven by changing characteristics of AgNPs with varying

48

environmental conditions. Investigation on a case-by-case basis may therefore be necessary.

49

Some studies have examined how various Ag NP coatings and surface charge influence their fate.

50

Tejamaya et al. observed higher aggregation of citrate-coated AgNPs in ecotoxicological test

51

media than polyethylene glycol (PEG)-coated AgNPs.13 Additionally, some studies have

52

revealed that sterically stabilized AgNPs are less affected by electrolyte composition, pH and

53

ionic strength compared to electrostatically stabilized AgNPs.14-17 The surface charge and

54

electrostatical interactions play also an important role in the ecotoxicological impact of

55

nanoparticles. El Badawy et al. reported higher toxicity of positively charged AgNPs to bacteria

56

compared to the negatively charged AgNPs and explained this effect by electrostatic attraction

57

between positively charged AgNPs and bacteria with negative surface charge.18


3


Page 4 of 32

58

Natural organic matter (NOM) is also an important factor influencing the fate and

59

ecotoxicity of AgNPs. Various types of NOM affect the colloidal stability and dissolution of

60

nanoparticles, with consequent changes in their ecotoxicity.7,19,20 Gunsolus et al. reported

61

increasing stabilization of citrate-coated AgNPs in the presence of sulfur- and nitrogen-rich

62

NOM.21 Collin et al. observed an increasing release of dissolved Ag even from sulfidized AgNPs

63

in the presence of Pony Lake Fulvic Acid (PLFA).22 Yang et al. showed that NOM, especially

64

PLFA, reduced the ecotoxicity of citrate-coated AgNPs.23 The previous studies provide a general

65

overview on the role of NOM in the fate and toxicity of AgNPs. Systematic studies are still

66

needed to verify the effects of NOM as well as oppositely charged capping agents of AgNPs on

67

their fate and ecotoxicity.

68

In order to provide a systematic and exhaustive understanding, in our study, we addressed

69

the special case of positively charged polymer (BPEI-) coated AgNPs with considering NOM

70

and their concentration to identify the detailed mechanisms controlling the aggregation,

71

dissolution, and ecotoxicological impact of AgNPs coated by oppositely charged organic

72

molecules or polymers in the absence and presence of NOM. In addition, we note the role of

73

formation of dissolved complexes between released Ag+ ions and NOM in the dissolution of

74

AgNPs and their toxicity. Pony Lake Fulvic Acid (PLFA), which has a higher content of sulfur

75

among commercially available reference standards of NOM, was selected as a model NOM.24,25

76

It is known from the literature, that the thiol ligands show high affinity to metallic Ag26 and Ag+

77

ions27. Regarding ecotoxicological assessment of AgNPs, Daphnia magna, a sensitive aquatic

78

species recommended by the Organization for Economic Cooperation and Development (OECD),

79

was used. There are several studies demonstrating the aggregation of AgNPs in ecotoxicological

80

test media.28,29 Therefore, a solution of the monovalent electrolyte NaNO3 was used at a low


4

Page 5 of 32


81

concentration of 10 mM as an acute toxicity exposure medium and as a negative control for all

82

exposure experiments to avoid possible changes in the colloidal stability of AgNPs during the

83

ecotoxicological tests.

84 85

EXPERIMENTAL SECTION

86

Materials

87

Two types of spherical silver nanoparticles (AgNPs) coated by citrate (Cit-) and branched

88

polyethylenimine (BPEI-), with an average core size of 23.5 ± 2.5 and 22.1 ± 1.2 nm,

89

respectively, were obtained from Nanocomposix, USA. Stock dispersions of both Cit-AgNPs and

90

BPEI-AgNPs were provided at the concentrations of 1.06 ± 0.01 mg/mL (in 2 mM of aqueous

91

citrate solution) and 1.00 ± 0.01 mg/mL (in Milli-Q water) as total silver (Ag), respectively. The

92

physicochemical properties of these nanoparticles are listed in Table S1.

93

Pony Lake Fulvic Acid (PLFA) was obtained from the International Humic Substances

94

Society (IHSS, USA), and its elemental composition is listed in Table S2.24 Details about the

95

preparation of stock and working solutions of PLFA are in the Supporting Information (SI).

96 97

Methods

98

Aggregation and dissolution of AgNPs

99

Aggregation and dissolution of Cit-AgNPs and BPEI-AgNPs were monitored in 10 mM

100

NaNO3 in both the absence and presence of PLFA. PLFA concentrations of 5 and 30 mg/L in the


5


Page 6 of 32

101

samples were within the range representative for the concentrations of NOM in natural

102

waters.30,31 The background concentration of NaNO3 in all samples maintained a typical ionic

103

strength for freshwater.32 To monitor the aggregation, dispersions of AgNPs were prepared at a

104

fixed concentration of 2 mg/L in both the absence and presence of PLFA in triplicate by diluting

105

the stock dispersions of AgNPs in the respective medium, followed by rotation at 35 rpm

106

(Rotator Genie, Scientific Industries, USA) in the dark and at room temperature. The above

107

samples were prepared in siliconized microcentrifuge tubes (Sigma-Aldrich, USA) to avoid

108

adsorption of nanoparticles onto the surface of the tubes. The aggregation of AgNPs was

109

assessed for 0, 1, 2, and 3 days using a Zetasizer Nano ZS (Malvern, Germany) and a UV-Vis

110

spectrometer (Perkin-Elmer, USA) from 235 to 700 nm. Hydrodynamic diameter (Dh) and

111

polydispersity index (PdI) were determined at 25°C by dynamic light scattering (DLS) using the

112

Zetasizer Nano ZS with a 173° back-scattering angle. Zeta potential values of AgNPs were also

113

determined using the same instrument and a disposable folded capillary cell (DTS1070, Malvern,

114

Germany) by applying the laser Doppler electrophoresis method in combination with phase

115

analysis light scattering (PALS) and multi-frequency measurements. To characterize the

116

morphology and validate the elemental composition of AgNPs in the presence of PLFA,

117

Scanning Electron Microscopy (SEM, Quanta 250 FEG, FEI Company, Holland) was used with

118

Energy Dispersive X-ray Spectroscopy (EDX). Experimental details for SEM are reported in SI.

119

Time-dependent dissolution of AgNPs was monitored in the absence and presence of

120

PLFA for 6 days covering 3 days of the pre-aging step and 48 hours of the exposition duration in

121

acute toxicity tests (details see in the next chapter). Quantification of dissolved Ag released from

122

AgNPs for both cases was achieved using an inductively coupled plasma mass spectrometer

123

(iCAP Qc ICP-MS, Thermo Fisher Scientific, Germany) and experimental details of the ICP-MS


6

Page 7 of 32


124

analysis are reported in SI. The dispersions of AgNPs were prepared at a fixed concentration of

125

100 µg/L in the absence and presence of PLFA in triplicate in silanized glass vials (Sigma-

126

Aldrich, USA) followed by storage in the dark at room temperature, rotating at 35 rpm. To

127

separate AgNPs from the dissolved fraction, the collected samples were centrifuged each day for

128

10 min at 4000 g (Heraeus Multifuge 1S-R, Thermo Scientific, Germany) using an ultrafiltration

129

membrane with a molecular weight cutoff (MWCO) of 50 kDa (Amicon Ultra-4 Centrifugal

130

Filter, Merck Millipore, Germany). Additionally, the ultrafiltration membrane with a MWCO of

131

3 kDa was also used to determine whether the released Ag form dissolved complexes with PLFA.

132

These samples were only monitored for the longest time (6 day) in triplicate using the same

133

method used for the samples treated with an ultrafiltration membrane with a MWCO of 50 kDa.

134

The samples were centrifuged for 30 min at 4000 g. The filtrate fractions from all samples were

135

acidified with concentrated nitric acid (TraceSelect® Ultra, Sigma-Aldrich, USA) to a final

136

concentration of 1.2% HNO3 prior to ICP-MS analysis.

137

Daphnia magna acute toxicity test

138

A 48-h acute toxicity test on Daphnia magna was performed using Daphtoxkit FTM

139

magna (MicroBioTests, Inc., Belgium) in quadruplicate, which follows OECD guideline 202.33

140

All tests were conducted for Cit-AgNPs and BPEI-AgNPs in the absence and presence of PLFA,

141

as well as for coating materials only, namely, sodium citrate tribasic dihydrate and branched

142

polyethylenimine with an average molecular weight (Mw) of 25000 Da (Sigma-Aldrich, USA).

143

Then, 10 mM NaNO3 was used as a medium for the acute toxicity test with Daphnia magna.

144

Exposure of all samples for AgNPs with and without PLFA was set at a concentration of 100

145

µg/L total Ag in the medium. According to the literature data, the predicted concentrations of

146

engineered AgNPs in surface waters are in the range from ca. 0.01 ng/L to ca. 1 µg/L.34 The


7


Page 8 of 32

147

concentration of AgNPs (100 µg/L) applied in our work is above but still close to the

148

environmentally relevant concentration range. According to other studies, the median effective

149

concentrations (48 h EC50) of AgNPs for Daphnia magna are in the range from 0.75 to 187 µg/L

150

(Seitz et al., 2015). In order to evaluate the impact of the dissolution of AgNPs on their adverse

151

effects to Daphnia magna, we used 100 µg/L of AgNPs, which is in the range of EC50 obtained

152

in other ecotoxicological studies for the same organism. Samples in the absence of PLFA were

153

directly exposed to Daphnia magna, but samples in the presence of PLFA were pre-aged by

154

shaking on an orbital shaker at 135 rpm for 3 days in the dark and at room temperature prior to

155

the exposure to Daphnia magna. The pre-aging step of AgNPs samples in the presence of PLFA

156

was necessary to minimize or avoid changes in nanoparticle transformation and colloidal state

157

during the ecotoxicological test with Daphnia magna. Ten Daphnia magna (less than 24 h old)

158

per treatment were exposed to each sample mentioned above at pH 7 and 22.5 ± 0.5°C with a 16

159

h:8 h (light:dark) photoperiod cycle. Adverse effects (%), represented as amount of ecotoxicity,

160

were calculated as the percentage of immobilization, including mortality of Daphnia magna in

161

the treatment group.35 The content of total Ag, including dissolved Ag and AgNPs in the entire

162

Daphnia magna body was quantified by ICP-MS after a sample preparation step; detailed

163

information is given in the SI.

164 165

RESULTS AND DISCUSSION

166

Impact of PLFA on colloidal stability of AgNPs

167

The hydrodynamic diameter (Dh) of Cit-AgNPs in Milli-Q water as measured by DLS

168

was 23.9 ± 0.1 nm (Table S1), in the same range as the core diameter (23.5 ± 2.5 nm) determined


8

Page 9 of 32


169

by SEM measurements. In contrast, for BPEI-AgNPs, a higher Dh of 64.6 ± 5.3 nm was

170

determined in Milli-Q water compared to the core diameter (22.1 ± 1.2 nm). This indicates that

171

BPEI-AgNPs were slightly aggregated in Milli-Q water. As reported by El Badawy et al. (2010),

172

the BPEI-AgNPs are aggregated at pH values below 6 and above 7 in a solution of 10 mM

173

NaNO3 by bridging between BPEI coatings.14 At neutral pH values (6-7), the structural

174

configuration of BPEI can be changed and the molecules form more rigid structures due to

175

hydrogen bonding between free and charged amine groups.14,36 These conformational changes of

176

BPEI lead to stabilization of BPEI-AgNPs, predominantly through steric interactions.14 The pH

177

value measured in our initial dispersion of BPEI-AgNPs was 7.04 and thereby on the upper limit

178

of the pH range (6-7) where the colloidal stability of BPEI-AgNPs is highest. Therefore, slight

179

changes in pH which cannot be excluded can result in aggregation of BPEI-AgNPs. On the other

180

hand, the adsorption of BPEI molecules with molecular weight of 25 kDa will result in

181

increasing Dh of AgNPs. While Dh measured by DLS usually comprise the core size with the

182

coating polymer layer, measurement of AgNPs using SEM only provides core size of AgNPs if

183

no special methods are applied to visualize the polymer coating. As already reported in the

184

literature, particles coated by polymers, depending on their molecular mass, show a significantly

185

larger Dh compared to the core size.37,38 Slightly higher aggregation of Cit-AgNPs compared to

186

BPEI-AgNPs observed on the SEM images (Table S1) is most probably an effect of sample

187

preparation. Since the Cit-AgNPs and BPEI-AgNPs in stock dispersions are dispersed in liquid

188

phases with different chemistry (solution of 2 mM citrate and Milli-Q water, respectively), the

189

drying during sample preparation can result in different degree of sample changes.

190

Dh and PdI values of AgNPs were also determined in 10 mM NaNO3 solution in the

191

absence and presence of PLFA using DLS (Figure 1, Table S4). The initial sample (0 day in


9


Page 10 of 32

192

Figure 1 and Table S4), should be considered as 6 min, the time gap between sample preparation

193

and measurement. After the first 6 min, the Cit-AgNPs and BPEI-AgNPs dispersed in 10 mM

194

NaNO3 solution in the absence of PLFA showed a mean Dh of 52.1 ± 3.0 and 42.5 ± 0.4 nm, and

195

a PdI of 0.11 and 0.25, respectively, indicating slight aggregation and a low or moderate width of

196

the size distribution. PdI can be used to describe the broadness of the particle size distribution.

197

Values below 0.1 indicate highly homogeneous particle size, whereas higher values especially

198

close to 1 indicate a broad size distribution and the presence of large particles or aggregates.39,40

199

In the absence of PLFA Cit-AgNPs and BPEI-AgNPs also showed differences in aggregation

200

during the 3 day period. While after 3 days BPEI-AgNPs showed low aggregation (Dh = 157.4 ±

201

131.1 nm) and a moderate size distribution (PdI = 0.30), Cit-AgNPs revealed strong aggregation

202

(Dh = 3127.5 ± 193.0 nm) and a broad size distribution (PdI = 1.00). Consequently, in the

203

absence of PLFA, the colloidal stability of BPEI-Ag NP was higher than Cit-AgNPs.

204

After exposure to PLFA, aggregation differed between the AgNPs. Cit-AgNPs exposed to

205

5 and 30 mg/L of PLFA showed high colloidal stability and a moderate particle size distribution

206

over 3 days of incubation, as demonstrated by a Dh of 56.4-62.0 nm and 50.0-81.8 nm, with an

207

average PdI of 0.33-0.36 and 0.18-0.31, respectively (Table S4). Cit-AgNPs showed similar

208

aggregation dynamics at both concentrations of PLFA (Figure 1 (A)). In contrast, BPEI-AgNPs

209

were immediately aggregated in the presence of PLFA and settled with large particles of 4880.0

210

± 3355.9 and 1456.0 ± 329.5 nm and PdI values close to 1.00 (Table S4). For 3 days of

211

incubation with PLFA, the aggregation of BPEI-AgNPs differed depending on the concentrations

212

of PLFA (Figure 1 (C)). BPEI-AgNPs exposed to 5 mg/L of PLFA showed strong aggregation,

213

with particle sizes of 1373.0-4880.0 nm and PdI values near 1.00; the extent of aggregation and

214

the PdI of BPEI-AgNPs decreased in the presence of 30 mg/L PLFA, as demonstrated by the Dh


10

Page 11 of 32


215

of 181.7-1456.0 nm and PdI of 0.33-0.92. This suggests that the increase in PLFA concentration

216

enhances colloidal stability and homogeneity in particle size for BPEI-AgNPs. The presence of

217

low concentration of PLFA resulted in the adsorption of negatively charged PLFA to the

218

positively charged surface of BPEI-AgNPs leading to charge screening and aggregation of

219

nanoparticles. At higher concentration further adsorption of PLFA enhanced the colloidal

220

stability and dispersion of nanoparticles due to the electrostatic and steric hindrance. Decreased

221

aggregation of BPEI-AgNPs at a high PLFA concentration is consistent with previous research

222

by Palomino et al.41 They observed that aggregation of hematite nanoparticles is promoted by the

223

adsorption of fulvic acid on the surface of positively charged nanoparticles, but disaggregation of

224

already aggregated nanoparticles is led by increasing fulvic acid concentration.

225

The zeta potentials of AgNPs in the absence and presence of PLFA are shown in Figure 1

226

and Table S5. Cit-AgNPs were charged negatively in the absence (zeta potentials from -13.9 to -

227

21.6 mV) and presence (zeta potentials from -16.5 to -23.0 mV) of PLFA. In contrast, the charge

228

of BPEI-AgNPs was positive in the absence of PLFA (zeta potentials from 12.2 to 21.4 mV) and

229

shifted to negative values only in the presence of PLFA (zeta potentials from -17.8 to -25.3 mV).

230

Differences in the aggregation of Cit-AgNPs and BPEI-AgNPs may be induced by differences in

231

surface charge and stabilization mechanisms. Negatively charged Cit-AgNPs were only

232

electrostatically stabilized by deprotonated carboxyl groups of citrate which is weakly bound to

233

the surface of AgNPs; the citrate may be readily displaced by PLFA which has high affinity to

234

metallic and ionic silver.14 In addition, replacement of citrate by PLFA may contribute to more

235

effective colloidal stability of Cit-AgNPs through greater steric repulsion by PLFA. In contrast,

236

BPEI, more strongly bound on the surface of AgNPs than citrate, may be harder to displace.

237

BPEI-AgNPs were strongly stabilized by electrostatic repulsion and steric repulsion through


11


Page 12 of 32

238

polymer chains of BPEI that provide a positive charge on the surface of AgNPs due to their

239

amino groups.42 Electrostatic interactions between negatively charged PLFA and positively

240

charged BPEI-AgNPs may play an important role in aggregation. The interaction of PLFA with

241

BPEI coating may result in the adsorption of PLFA to the surface of BPEI-AgNPs, leading to

242

charge screening and enhanced aggregation of particles. At neutral pH, the phenolic and

243

carboxylic functional groups of fulvic acids are usually deprotonated, and the charge of PLFA

244

becomes negative.43,44 Therefore, in the presence of PLFA, BPEI-AgNPs showed charge

245

inversion from positive to negative.

246


12

Page 13 of 32


3500

(A)

(B)

-5

Zeta potential (mV)

Dh (nm)

3250

0

Cit-AgNPs in the absence of PLFA Cit-AgNPs in 5 mg/L PLFA Cit-AgNPs in 30 mg/L PLFA

3000 200 150 100 50 0

Cit-AgNPs in the absence of PLFA Cit-AgNPs in 5 mg/L PLFA Cit-AgNPs in 30 mg/L PLFA

-10 -15 -20 -25 -30

0

1

2

3

0

1

Time (day) 10000 7500

(D)

BPEI-AgNPs in the absence of PLFA BPEI-AgNPs in 5 mg/L PLFA BPEI-AgNPs in 30 mg/L PLFA

40

Zeta potential (mV)

5000

Dh (nm)

2500 250 200 150 100 50 0 0

1

3

Time (day) 60

(C)

2

2

3

BPEI-AgNPs in the absence of PLFA BPEI-AgNPs in 5 mg/L PLFA BPEI-AgNPs in 30 mg/L PLFA

20 0 -20

0

1

2

3

Time (day)

Time (day)

247 248

Figure 1. Mean hydrodynamic diameters (Dh, nm) (A and C) and mean zeta potentials (mV) (B

249

and D) of 2 mg/L Cit-AgNPs and BPEI-AgNPs incubated in 10 mM NaNO3 solution in the

250

absence and presence of 5 and 30 mg/L PLFA for 3 days. The error bars represent standard

251

deviation of 3 replicates.

252

Figure S1 shows UV-vis absorption spectra of AgNPs incubated in the absence and

253

presence of 5 and 30 mg/L PLFA for 3 days. After 6 min, spherical Cit-AgNPs and BPEI-AgNPs

254

dispersed in 10 mM NaNO3 solution in the absence of PLFA showed surface plasmon resonance

255

(SPR) peaks at 402 and 407 nm, respectively. Over 3 days of incubation at both concentrations

256

of PLFA, as shown in Figure S1 (A) and (B), only a slight decrease in SPR peak intensity for


13


Page 14 of 32

257

Cit-AgNPs was observed near 400 nm, without broadening and a red-shift in SPR peaks,

258

indicating that Cit-AgNPs were stabilized in the presence of PLFA. These results are in good

259

agreement with DLS measurements, demonstrating high stability of Cit-AgNPs in the presence

260

of 5 and 30 mg/L PLFA. The slight decrease in absorbance intensity may be correlated to the

261

partial surface passivation of Cit-AgNPs by PLFA.45 In contrast, BPEI-AgNPs incubated with 5

262

mg/L of PLFA showed broadening and a red-shift, as well as a decrease in the intensity of SPR

263

peaks over 3 days of incubation (Figure S1 (C)). This is attributed to the aggregation and broad

264

size distribution of aggregates, which was also demonstrated by increasing Dh and high values of

265

PdI in the DLS measurements. The broadening and red-shifting of the SPR peak can be

266

explained by different charge displacements and multipolar charge distributions of aggregates.46

267

When the particles have a homogeneous size distribution and the size is much smaller than the

268

wavelength of incident light, the electron cloud of the particles shows a homogeneous dipole

269

charge distribution and only one plasmon resonance. However, the increasing particle size yields

270

high multipolar charge distributions with shifting the position of the dipolar mode to larger

271

wavelengths. In this case, the absorption spectrum shows red-shifted broader surface plasmon

272

resonances. BPEI-AgNPs incubated with 30 mg/L of PLFA also showed peak broadening and

273

red-shifting near 400 nm (Figure S1 (D)). In contrast to the decreasing absorbance for BPEI-

274

AgNPs in the presence of 5 mg/L of PLFA, BPEI-AgNPs over 3 days of incubation with 30

275

mg/L of PLFA showed fluctuating SPR peak intensities, as observed in the absorbance from 0.12

276

to 0.15 A.U. near 430 nm between 0 and 1 day of incubation, followed by a slight but continuous

277

decrease in absorbance with increasing incubation time.

278

The results from DLS and UV-vis spectroscopy demonstrated that PLFA strongly effects

279

the dispersion stability of both Cit-AgNPs and BPEI-AgNPs. PLFA, at both applied


14

Page 15 of 32


280

concentrations (5 and 30 mg/L), enhanced the colloidal stability of Cit-AgNPs. The stability of

281

BPEI-AgNPs, however, was only partly enhanced at a PLFA concentration of 30 mg/L

282

compared to the PLFA concentration of 5 mg/L.

283

The results of SEM analysis of BPEI-AgNPs in Milli-Q water as well as in the presence

284

of PLFA are shown in Figure 2. BPEI-AgNPs in Milli-Q water without PLFA exhibited

285

nanosized particles, while in the presence of PLFA much larger structures were observed. EDX

286

analyses at positions a and b were also carried for the BPEI-Ag NP sample in the presence of

287

PLFA (Figure 2 (B)). As listed in Table 1, S and Ag were determined at position a. No signal for

288

S and Ag, however, was observed for position b. Furthermore, the signal for C was significantly

289

higher in position a compared to the background (position b). The appearance of a low signal for

290

C in the background can be related to the fraction of PLFA remaining in the solution (not bound

291

to the AgNPs). For the formation of the large structures shown in Figure 2 (B), the influence of

292

drying effects during sample preparation cannot be fully excluded, but the results of EDX

293

measurements indicate that the large structures most likely represent the aggregates of BPEI-

294

AgNPs embedded in a PLFA matrix. In addition, it is known that the S and N present in NOM

295

are representative elements with high affinity to metallic and ionic silver.25,47 Therefore, we note

296

that they can significantly contribute to the formation of Ag NP aggregates with fulvic acid.


15


Page 16 of 32

(B)

(A)

a

b 500 nm

400 nm

297 298

Figure 2. SEM images of 10 mg/L BPEI-AgNPs in Milli-Q water (A) and 1 mg/L BPEI-AgNPs

299

in the presence of 15 mg/L PLFA (B). Energy Dispersive X-ray (EDX) analysis was carried at

300

the positions of a and b in Figure (B) for the BPEI-AgNPs sample incubated in the presence of

301

PLFA. The elemental compositions considered in this study are listed in Table 1.

302 303

Table 1. Elemental compositions detected in the sample of BPEI-AgNPs incubated with 15

304

mg/L PLFA, corresponding to Figure 2 (B). a

b

Weight %

Atomic %

Weight %

Atomic %

C

47.2

75.5

12.3

25.4

Ag

8.4

1.5

-

-

S

1.6

0.9

-

-

Si

30.5

20.9

82.7

73.2

Au

12.3

1.2

5.0

1.4

305 306


16

Page 17 of 32

307


Impact of PLFA on dissolution of AgNPs

308

The concentration of dissolved Ag released from Cit-AgNPs was higher than that of

309

BPEI-AgNPs in 10 mM NaNO3 in the absence of PLFA (Figure 3 (A)). For samples labeled as 0

310

day, an initial preparation time of approximately 10 min was used. After 6 days, the

311

concentrations of dissolved Ag released from Cit-AgNPs and BPEI-AgNPs were 7.5 ± 1.6 and

312

4.1 ± 0.6 µg/L, respectively (as shown in Table S6). Cit-AgNPs showed approximately 2 times

313

more dissolved Ag than BPEI-AgNPs compared for the average concentrations for the 6 days of

314

the experimental period. This difference in the release of dissolved Ag may be attributed to the

315

number of active sites on the surface of AgNPs, which can be attacked by dissolved oxygen.48,49

316

AgNPs coated with citrate may be more prone to be attacked by dissolved oxygen, followed by

317

an oxidation process, than BPEI-AgNPs, which are sterically stabilized by polymer chains of

318

BPEI acting as a chemical barrier.50 This indicates that the accessibility of oxidation sites of

319

AgNPs may play an important role in determining the dissolution of AgNPs. The release of

320

dissolved Ag from AgNPs may also depend on different surface functional groups of coating

321

materials bound on the surface of AgNPs.51,52 The initial release of dissolved Ag from both

322

AgNPs may be hindered by the re-sorption of dissolved Ag to the AgNPs surface49,52 due to the

323

binding of dissolved Ag to carboxyl groups of citrate and amino groups of BPEI on the surface

324

of AgNPs. The binding ability of the functional groups also has an important role in releasing

325

dissolved Ag from surface-coated AgNPs. Nitrogen of the amino groups on BPEI-AgNPs, which

326

binds strongly to dissolved Ag compared to the carboxyl groups on Cit-AgNPs, may additionally

327

lead to less release of dissolved Ag from BPEI-AgNPs.25

328

The presence of PLFA influenced the dissolution of AgNPs. For both AgNPs, decrease in

329

the concentration of released dissolved Ag was observed in the presence of 5 mg/L PLFA. The


17


Page 18 of 32

330

reduced dissolution of AgNPs may be the result of the shielding of active sites of AgNPs by the

331

steric hindrance of PLFA.49 The presence of 30 mg/L PLFA, however, resulted in higher release

332

of dissolved Ag compared to experiments performed at 5 mg/L PLFA for both AgNPs and in the

333

absence of PLFA only for BPEI-AgNPs. For Cit-AgNPs no significant differences in the release

334

of dissolved Ag could be observed in the absence and presence of 30 mg/L PLFA. After 6 days

335

of incubation, the concentrations of dissolved Ag released from Cit-AgNPs and BPEI-AgNPs in

336

the presence of 30 mg/L PLFA were increased to 9.5 ± 0.4 and 7.1 ± 0.6 µg/L, respectively

337

(Table S6). This indicates that NOM, depending on concentration, may induce decreasing and

338

increasing dissolution of AgNPs. Several researchers observed that higher concentrations of

339

NOM inhibit the release of dissolved Ag.49,53 However, in this study, the concentration of

340

dissolved Ag released from AgNPs in the presence of 30 mg/L PLFA was higher than in the

341

presence of 5 mg/L PLFA, as shown in Figure 3 (B) and (C) and Table S6. The differences in

342

dissolution of AgNPs depending on the concentration of PLFA may be related to its sulfur and

343

nitrogen content. The increasing concentration of PLFA may lead to increased complexing of

344

released dissolved Ag with PLFA, which will result in increased release of dissolved Ag by the

345

shifting the chemical equilibrium of dissolution towards the dissolution of nanoparticles (i.e., Le

346

Châtelier’s principle). Enhanced dissolution of both AgNPs at high concentrations of PLFA

347

observed in our study is in good agreement with the studies of Yang et al. and Collin et al.22,23,

348

who reported increased release of Ag from citrate-coated AgNPs with increasing concentration

349

of PLFA and from sulfidized AgNPs in the presence of PLFA, respectively. In order to

350

distinguish between free Ag+ ions and Ag+ ions complexed by PLFA, the separation of AgNPs

351

from liquid phase in dissolution experiments with the duration of 6 day was done using the

352

ultrafiltration membrane with a MWCO of 3 kDa. As shown in Table S7, the concentrations of


18

Page 19 of 32


353

dissolved Ag after filtration using the membrane with a MWCO of 3 kDa were lower than for a

354

MWCO of 50 kDa. This means that the released dissolved Ag likely forms dissolved complexes

355

with PLFA. The differences between concentrations of dissolved Ag in 3 kDa and 50 kDa

356

filtrates increased with increasing concentration of PLFA. In 3 kDa filtrate the concentrations of

357

dissolved Ag in the presence of 30 mg/L PLFA was even lower than the 0.1 µg/L, the limit of

358

quantification (LOQ) of ICP-MS. Based on these results, we suggest that the dissolved Ag

359

released from AgNPs in the presence of PLFA is predominantly complexed by PLFA fractions,

360

with molecular weights between 3 kDa and 50 kDa. Only a small portion of the total released Ag

361

was available as free Ag+ ions or dissolved complexes with the PLFA fraction < 3 kDa. The

362

formation of dissolved complexes between dissolved Ag and PLFA can shift the dissolution

363

equilibrium of AgNPs, followed by enhancing the concentration of released Ag. Furthermore,

364

the role of partial stabilization of BPEI-AgNPs observed by DLS measurements in the presence

365

of 30 mg/L PLFA and inevitably associated with increasing surface area, which can also lead to

366

the increasing release of dissolved Ag, cannot be fully neglected.

367

Similar to previous studies by other researchers,48-50,52,53 our results suggest that the

368

accessibility of oxidation sites of AgNPs, concentration of NOM, and binding ability of their

369

functional groups with dissolved Ag might play important roles in determining the dissolution of

370

AgNPs in the presence of NOM. The speciation and concentration of Ag released from AgNPs

371

most likely depend on the concentration of NOM and their ability to form dissolved complexes

372

with dissolved Ag.22,23,25,49 Since the type of dissolved Ag species released from AgNPs and their

373

bioavailability are key factors influencing the uptake of Ag by organisms, the findings of this

374

study may contribute to a better understanding of the adverse effects of AgNPs in the presence of

375

NOM.


19


Ag concentration (µg/L)

14

(A)

Cit-AgNPs in the absence of PLFA BPEI-AgNPs in the absence of PLFA

14

(B)

Cit-AgNPs in 5 mg/L PLFA Cit-AgNPs in 30 mg/L PLFA

Page 20 of 32

14

12

12

12

10

10

10

8

8

8

6

6

6

4

4

4

2

2

2

0

0

0

1

2

3

Time (day)

4

5

6

(C)

BPEI-AgNPs in 5 mg/L PLFA BPEI-AgNPs in 30 mg/L PLFA

0 0

1

2

3

4

5

6

0

1

Time (day)

2

3

4

5

6

Time (day)

376

Figure 3. Mean concentrations (µg/L) of dissolved Ag released from 100 µg/L of Cit-AgNPs

377

and BPEI-AgNPs dispersed in 10 mM NaNO3 solution in the absence (A) and presence (B and

378

C) of 5 and 30 mg/L PLFA for 6 days. The error bars represent standard deviation of 3 replicates.

379 380

Impact of PLFA on ecotoxicological effects of AgNPs to Daphnia magna

381

The calculated adverse effects (%) are presented in Figure 4 for AgNPs dispersed in 10

382

mM NaNO3 solution in both the absence and presence of PLFA, as well as for coating materials

383

only. The adverse effects of the 10 mM NaNO3 solution, used as an ecotoxicity exposure

384

medium and as a negative control, was shown as 10.4 ± 1.6%. The test condition was valid

385

because the mortality in the control was not higher than 10% (≤10%).33 The adverse effect of the

386

citrate coating material alone, without nanoparticles, was not significantly different from that of

387

Cit-AgNPs in the presence of PLFA. However, among all samples tested in the study, the BPEI

388

coating material alone exhibited the highest toxic effect to Daphnia magna which may be

389

attributed to the surface charge interaction between positively charged BPEI and the negatively

390

charged cell membrane of Daphnia magna at pH 6-7.54 In the absence of PLFA, Cit-AgNPs

391

showed higher ecotoxicological effects than BPEI-AgNPs, but the adverse effects of both AgNPs

392

were reduced after exposure to PLFA. The adverse effects of Cit-AgNPs and BPEI-AgNPs in the


20

Page 21 of 32


393

presence of PLFA were approximately 70% and 60% lower than in the absence of PLFA,

394

respectively. These adverse effects of AgNPs may be attributed to the concentration of dissolved

395

Ag, as suggested by Pokhrel et al.5 and Lee et al.55. Seitz et al. also explained the reducing

396

ecotoxicological effects for AgNPs in the presence of dissolved organic matter (DOM) as a result

397

of blocking the active site on the surface of AgNPs for the release of Ag+ ions.7 According to

398

previous results reported for toxicity of soluble Ag, the median effective concentrations (48 h

399

EC50) of free dissolved Ag for Daphnia magna are in the range from 0.4 to 15.9 µg/L.55,56 In this

400

study, the concentrations of dissolved Ag released from Cit-AgNPs and BPEI-AgNPs in the

401

absence of PLFA were 4.4 ± 0.8 and 2.3 ± 0.7 µg/L, respectively (Table S6), corresponding to

402

concentrations exposed to Daphnia magna at 48 h. These numbers are in the range of the

403

previously reported EC50 and consequently indicate that the adverse effects of AgNPs in the

404

absence of PLFA may largely depend on the concentrations of free dissolved Ag. Some studies

405

showing the effects of coating materials on the toxicity of AgNPs identified less toxic effects of

406

coated AgNPs than uncoated AgNPs.7,57 However, we identified that in the absence of PLFA, the

407

release of free dissolved Ag is much more important to induce the toxicity than other properties

408

like size, aggregation state or original coating. Figure 4 shows the correlation between the

409

dissolved free Ag+ ions and adverse effects of AgNPs. The presence of free Ag+ ions can be

410

assumed in the samples without PLFA since the presence of free dissolved coating materials,

411

which can complex Ag+ ions, is not expected in the solution. Much lower adverse effects

412

determined in the absence of PLFA for BPEI-AgNPs compared to Cit-AgNP dispersion which

413

does not contain BPEI clearly show that the free BPEI if it is present in the dispersions of BPEI-

414

AgNPs only plays a subordinate role in the toxicity. However, to understand the correlation of

415

adverse effects and the dissolution of AgNPs in the presence of PLFA, we should consider that


21


Page 22 of 32

416

the concentrations of dissolved Ag measured in this study predominantly represent dissolved Ag

417

complexed with dissolved PLFA fractions, which are smaller than 50 kDa. In addition, the

418

presence of low concentrations of free Ag+ ions cannot be fully excluded.

419

The concentrations of dissolved Ag released from Cit-AgNPs and BPEI-AgNPs in the

420

presence of 5 mg/L PLFA at 5 days, which are representative for the Ag exposure concentrations

421

in Daphnia magna toxicity tests at 48 h, were 3.7 ± 0.3 and 2.3 ± 0.1 µg/L, respectively (Table

422

S6). This is not significantly different from the concentrations of dissolved Ag released from Cit-

423

AgNPs and BPEI-AgNPs at 2 days in the absence of PLFA, respectively. Although there is no

424

statistically significant difference in the concentration of dissolved Ag between them, both

425

AgNPs in the presence of 5 mg/L PLFA showed lower adverse effects compared to AgNPs in the

426

absence of PLFA. For Cit-AgNPs, the concentrations of dissolved Ag, representative for

427

Daphnia magna toxicity tests at 48 h, were higher at the higher concentration of PLFA (30

428

mg/L) than the values for 5 mg/L PLFA and in their absence, but the adverse effects of Cit-

429

AgNPs in the presence of 30 mg/L PLFA were not significantly different from that at 5 mg/L

430

PLFA and even lower than for the absence of PLFA. This result can be attributed to differences

431

in the concentrations of released free Ag+ ions. The amount of free dissolved Ag+ ions released

432

from AgNPs in the presence of PLFA might be lower than their concentration in the absence of

433

PLFA, due to complexation of free dissolved Ag with PLFA fractions. As shown in Table S7, in

434

the presence of PLFA, the released Ag was predominantly available as dissolved complexes with

435

PLFA. In the case of BPEI-AgNPs, the adverse effects in the presence of 5 mg/L and 30 mg/L

436

PLFA were also not significantly different. However, similar to Cit-AgNPs, higher adverse

437

effects of BPEI-AgNPs in the absence of PLFA were observed compared to the presence of

438

PLFA, which may be attributed to the higher concentration of free Ag+ ions released from


22

Page 23 of 32


439

AgNPs and the surface charge of BPEI-AgNPs. Positively charged BPEI-AgNPs can be adhesive

440

to the negatively charged cell membrane of Daphnia magna54,58 probably contributing partially

441

to the higher adverse effects of BPEI-AgNPs in the absence of PLFA than in their presence.

442

High toxicity of positively charged nanoparticles was reported in other studies for bacteria18 and

443

Daphnia magna59. However, in our study, the negatively charged Cit-AgNPs showed higher

444

adverse effects on Daphnia magna than positively charged BPEI-AgNPs in the absence of PLFA.

445

Most likely, the adverse effects of AgNPs are driven amongst others by concentration of released

446

toxic ions, as well as partially by electrostatic interactions between nanoparticles and cell

447

membranes of organisms. The content of total Ag such as dissolved Ag and AgNPs in Daphnia

448

magna may also have an important role in determining the ecotoxicological effects of BPEI-

449

AgNPs. The total Ag content in Daphnia magna (ng/µg dry weight) is presented in Table 2. The

450

total Ag mass fraction (%) in Daphnia magna was also calculated based on the total Ag mass in

451

10 of Daphnia magna (ng) per total Ag mass in samples of AgNPs (ng) that were previously

452

exposed to Daphnia magna. As listed in Table 2, the relatively higher Ag content of 1.4 ± 0.4

453

ng/µg dry weight (total Ag mass fraction of 15.8 ± 4.9%) was found in Daphnia magna exposed

454

to BPEI-AgNPs in the absence of PLFA, compared to Ag contents of 0.1 ± 0.0 and 0.3 ± 0.1

455

ng/µg dry weight (total Ag mass fraction of 1.3 ± 0.4% and 3.4 ± 1.0%) in the presence of both 5

456

and 30 mg/L PLFA, respectively. This suggests that high adverse effects of BPEI-AgNPs

457

without PLFA may also be attributed to the total Ag content in Daphnia magna. High content of

458

Ag in Daphnia magna in the absence of PLFA can be a result of the accumulation of positively

459

charged BPEI-AgNPs to the negatively charged cell membrane of Daphnia magna. However, the

460

content of total Ag in Daphnia magna exposed to Cit-AgNPs in the absence and presence of

461

PLFA were not different. Additionally, the Ag content in Daphnia magna exposed to Cit-AgNPs


23


Page 24 of 32

462

was comparable or lower than for BPEI-AgNPs. It is not clear that the total Ag content in

463

Daphnia magna exposed to Cit-AgNPs samples was from released Ag+ ions or AgNPs, but it

464

suggests that there might be limited effects of the total Ag in Daphnia magna on the

465

ecotoxicological potential of Cit-AgNPs. Furthermore, no considerable correlation was observed

466

between the colloidal state of AgNPs and their adverse effect. Most likely, the aggregation of

467

AgNPs plays only a subordinated role in their ecotoxicological impact on Daphnia magna in the

468

absence and presence of PLFA. The release of toxic Ag+ ions seems to be more responsible for

469

the adverse effects of AgNPs on Daphnia magna. Several studies also showed that free Ag+ ions

470

may play an important role in the determination of the ecotoxicological effects of AgNPs on

471

aquatic organisms.6,9-11 Although our results do not directly demonstrate the impact of free

472

dissolved Ag on the ecotoxicological effects of AgNPs to Daphnia magna, we showed that the

473

formation of dissolved complexes between released Ag+ ions and NOM as well as concentration

474

of NOM control the dissolution of AgNPs and their toxicity. We demonstrated that high

475

concentration of NOM enhances the release of dissolved Ag from AgNPs, but the increased

476

extent of AgNPs dissolution does not lead to the increasing toxicity since the major part of

477

released Ag was complexed with NOM.

478

The findings of this study demonstrate that the fate and ecotoxicological impact of

479

nanoparticles are determined by the interplay of several factors, like the type of nanoparticle

480

coating agents, release of toxic ions from nanoparticles, speciation of the released ions, presence

481

and concentration of NOM, as well as the surface charge of nanoparticles. High content of Ag in

482

Daphnia magna exposed to the positively charged AgNPs can be a risk for the accumulation and

483

enrichment of Ag in organisms feeding on Daphnia magna. Our results, therefore, are a sound

484

basis for further studies to understand the ecotoxicological impact of AgNPs particularly in the


24

Page 25 of 32


485

presence of NOM and to evaluate possible enrichment of positively charged nanoparticles in the

486

food web. Based on our results, we demonstrated the implications of dissolved Ag complexes

487

between Ag+ ions and NOM for the dissolution of AgNPs and for their toxicity. Especially, our

488

observation on the sensitivity of colloidal stability and dissolution of positively charged AgNPs

489

to the concentration of NOM indicates that their transport, mobility, and toxicity in aquatic

490

systems showing high degree of variation in chemical composition might depend on the location

491

and seasonal changes. The findings of our study will facilitate the prediction of the transport,

492

mobility and ecotoxicity of AgNPs depending on their surface charge and concentration of

493

NOM. Furthermore, our results, showing Ag NP aggregates associated with sulfur and embedded

494

in PLFA, as well as the complex relationship between surface properties of AgNPs, extent of

495

their dissolution, concentration of NOM and ecotoxicological effects of AgNPs suggest that

496

additional studies are required to understand in detail the complexity of the interactions and

497

toxicity mechanism for engineered nanoparticles released in natural environment.

498 499 500 501 502 503


25


100

Adverse effect (%) Ag concentration (µg/L)

Page 26 of 32

10 Cb

Adverse effects (%)

80 6 60

Ab

4 40 20

Ac, Ca

Ac Ac, Ca

--

Bb, Ca

Bb

2

Ca --

0

Ag concentration (µg/L)

8

Bc

0

I s 3 s te NO Citra gNP PLFA PLFA BPE AgNP PLFA PLFA Na A 0 5 0 5 I Cit Ps + s + 3 PE Ps + s + 3 mM B 0 P P N gN 1 gN N Ag -A Ag EI- EI-A Cit CitP B BP

504

Figure 4. The left y-axis: Mean adverse effects (%) of AgNPs (100 µg/L) dispersed in 10 mM

505

NaNO3 solution in the absence and presence of 5 and 30 mg/L PLFA, as well as coating

506

materials alone (100 µg/L), on Daphnia magna. Mean values and standard deviations were

507

calculated for 4 replicates. The t-test with 95% of confidential interval was performed to

508

compare mean adverse effects between AgNPs samples. Ac, Bc and Ab, Bb, Cb denote

509

significant differences for Cit-AgNPs and BPEI-AgNPs samples, respectively (p < 0.05). Ca

510

indicates a t-test result between control (10 mM NaNO3) and AgNPs samples. The same letters

511

in each AgNPs sample group denote no significant differences. The right y-axis: Mean

512

concentrations and standard deviations (µg/L) of dissolved Ag released from both AgNPs in the

513

absence (smaller than 50 kDa) as well as in the presence of PLFA (smaller than 3 kDa)

514

corresponding to the conditions at the end of Daphnia magna 48-h acute toxicity tests. Mean

515

concentrations and standard deviations were calculated for 3 replicates. The dissolved Ag

516

concentrations released from both AgNPs at high concentration of PLFA (30 mg/L) were below

517

the LOQ (limit of quantification) for ICP-MS, 0.1 µg/L for Ag.


26

Page 27 of 32


518

Table 2. Total Ag content (ng/µg dry weight) and total Ag mass fraction (%) in Daphnia magna

519

(mean values and standard deviations were calculated for 3 replicates). Samples exposed to Daphnia magna Cit-AgNPs (100 µg/L) Cit-AgNPs (100 µg/L) + PLFA (5 mg/L) Cit-AgNPs (100 µg/L) + PLFA (30 mg/L) BPEI-AgNPs (100 µg/L) BPEI-AgNPs (100 µg/L) + PLFA (5 mg/L) BPEI-AgNPs (100 µg/L) + PLFA (30 mg/L)

Total Ag content (ng/µg dry weight) 0.1 ± 0.0 0.1 ± 0.0 0.1 ± 0.0 1.4 ± 0.4 0.1 ± 0.0 0.3 ± 0.1

Total Ag mass fraction (%) 1.0 ± 0.3 1.4 ± 0.2 1.2 ± 0.2 15.8 ± 4.9 1.3 ± 0.4 3.4 ± 1.0

520 521 522

ASSOCIATED CONTENT

523

Supporting Information

524

Detailed experimental methods: Preparation of PLFA working solution, Scanning Electron

525

Microscopy (SEM), Total Ag content in Daphnia magna, ICP-MS measurements, and Statistical

526

analysis. Supporting figures and tables: UV-vis absorption spectra (Figure S1), Physicochemical

527

properties of AgNPs (Table S1), Elemental composition of PLFA (Table S2), Operating

528

parameters for ICP-MS (Table S3), Mean Dh and PdI of AgNPs in the absence and presence of

529

PLFA (Table S4), Mean zeta potential of AgNPs in the absence and presence of PLFA (Table

530

S5), and Mean concentrations of dissolved Ag released from AgNPs (Table S6 and Table S7).

531

This material is available free of charge via the Internet at http://pubs.acs.org.

532 533

AUTHOR INFORMATION

534

Corresponding Author

535

*Phone: +49 (0)681 9382 341; e-mail: [email protected]


27


Page 28 of 32

536

ACKNOWLEDGMENT

537

The authors are grateful to all the members of Environmental Safety Group in KIST Europe to

538

support us. This study was supported by KIST Europe Basic Fund Joint Research (Project #

539

11792) and the German Research Foundation (DFG, Research unit INTERNANO: FOR 1536,

540

subproject MASK SCHA849/16-2). The authors declare no competing financial interest.

541

REFERENCES

542 543 544 545 546 547 548 549 550 551 552 553 554 555 556 557 558 559 560 561 562 563 564 565 566 567 568 569 570 571 572 573 574

1. Benn, T. M.; Westerhoff, P., Nanoparticle silver released into water from commercially available sock fabrics. Environ. Sci. Technol. 2008, 42, (11), 4133-4139. 2. Duran, N.; Marcato, P. D.; Alves, O. L.; Da Silva, J. P. S.; De Souza, G. I. H.; Rodrigues, F. A.; Esposito, E., Ecosystem protection by effluent bioremediation: Silver nanoparticles impregnation in a textile fabrics process. J. Nanopart. Res. 2010, 12, (1), 285-292. 3. Limpiteeprakan, P.; Babel, S.; Lohwacharin, J.; Takizawa, S., Release of silver nanoparticles from fabrics during the course of sequential washing. Environ. Sci. Pollut. Res. 2016, 23, (22), 22810-22818. 4. McGillicuddy, E.; Murray, I.; Kavanagh, S.; Morrison, L.; Fogarty, A.; Cormican, M.; Dockery, P.; Prendergast, M.; Rowan, N.; Morris, D., Silver nanoparticles in the environment: Sources, detection and ecotoxicology. Sci. Total Environ. 2017, 575, 231-246. 5. Pokhrel, L. R.; Dubey, B.; Scheuerman, P. R., Impacts of select organic ligands on the colloidal stability, dissolution dynamics, and toxicity of silver nanoparticles. Environ. Sci. Technol. 2013, 47, (22), 12877-12885. 6. Newton, K. M.; Puppala, H. L.; Kitchens, C. L.; Colvin, V. L.; Klaine, S. J., Silver nanoparticle toxicity to Daphnia magna is a function of dissolved silver concentration. Environ. Toxicol. Chem. 2013, 32, (10), 2356-2364. 7. Seitz, F.; Rosenfeldt, R. R.; Storm, K.; Metreveli, G.; Schaumann, G. E.; Schulz, R.; Bundschuh, M., Effects of silver nanoparticle properties, media pH and dissolved organic matter on toxicity to Daphnia magna. Ecotox. Environ. Safe. 2015, 111, 263-270. 8. Ribeiro, F.; Van Gestel, C. A.; Pavlaki, M. D.; Azevedo, S.; Soares, A. M.; Loureiro, S., Bioaccumulation of silver in Daphnia magna: Waterborne and dietary exposure to nanoparticles and dissolved silver. Sci. Total Environ. 2017, 574, 1633-1639. 9. Wang, Z.; Chen, J.; Li, X.; Shao, J.; Peijnenburg, W. J., Aquatic toxicity of nanosilver colloids to different trophic organisms: Contributions of particles and free silver ion. Environ. Toxicol. Chem. 2012, 31, (10), 2408-2413. 10. Silva, T.; Pokhrel, L. R.; Dubey, B.; Tolaymat, T. M.; Maier, K. J.; Liu, X. F., Particle size, surface charge and concentration dependent ecotoxicity of three organo-coated silver nanoparticles: Comparison between general linear model-predicted and observed toxicity. Sci. Total Environ. 2014, 468, 968-976. 11. Navarro, E.; Piccapietra, F.; Wagner, B.; Marconi, F.; Kaegi, R.; Odzak, N.; Sigg, L.; Behra, R., Toxicity of Silver Nanoparticles to Chlamydomonas reinhardtii. Environ. Sci. Technol. 2008, 42, (23), 8959-8964.


28

Page 29 of 32

575 576 577 578 579 580 581 582 583 584 585 586 587 588 589 590 591 592 593 594 595 596 597 598 599 600 601 602 603 604 605 606 607 608 609 610 611 612 613 614 615 616 617 618 619


12. Taylor, C.; Matzke, M.; Kroll, A.; Read, D. S.; Svendsen, C.; Crossley, A., Toxic interactions of different silver forms with freshwater green algae and cyanobacteria and their effects on mechanistic endpoints and the production of extracellular polymeric substances. Environ. Sci. Nano 2016, 3, (2), 396-408. 13. Tejamaya, M.; Romer, I.; Merrifield, R. C.; Lead, J. R., Stability of citrate, PVP, and PEG coated silver nanoparticles in ecotoxicology media. Environ. Sci. Technol. 2012, 46, (13), 7011-7017. 14. El Badawy, A. M.; Luxton, T. P.; Silva, R. G.; Scheckel, K. G.; Suidan, M. T.; Tolaymat, T. M., Impact of environmental conditions (pH, ionic strength, and electrolyte type) on the surface charge and aggregation of silver nanoparticles suspensions. Environ. Sci. Technol. 2010, 44, (4), 1260-1266. 15. Hotze, E. M.; Phenrat, T.; Lowry, G. V., Nanoparticle aggregation: Challenges to understanding transport and reactivity in the environment. J. Environ. Qual. 2010, 39, (6), 19091924. 16. Huynh, K. A.; Chen, K. L., Aggregation kinetics of citrate and polyvinylpyrrolidone coated silver nanoparticles in monovalent and divalent electrolyte solutions. Environ. Sci. Technol. 2011, 45, (13), 5564-5571. 17. Ellis, L.-J. A.; Valsami-Jones, E.; Lead, J. R.; Baalousha, M., Impact of surface coating and environmental conditions on the fate and transport of silver nanoparticles in the aquatic environment. Sci. Total Environ. 2016, 568, 95-106. 18. El Badawy, A. M.; Silva, R. G.; Morris, B.; Scheckel, K. G.; Suidan, M. T.; Tolaymat, T. M., Surface charge-dependent toxicity of silver nanoparticles. Environ. Sci. Technol. 2011, 45, (1), 283-287. 19. Philippe, A.; Schaumann, G. E., Interactions of dissolved organic matter with natural and engineered inorganic colloids: A review. Environ. Sci. Technol. 2014, 48, (16), 8946-8962. 20. Aiken, G. R.; Hsu-Kim, H.; Ryan, J. N., Influence of dissolved organic matter on the environmental fate of metals, nanoparticles, and colloids. Environ. Sci. Technol. 2011, 45, (8), 3196-3201. 21. Gunsolus, I. L.; Mousavi, M. P. S.; Hussein, K.; Buhlmann, P.; Haynes, C. L., Effects of humic and fulvic acids on silver nanoparticle stability, dissolution, and toxicity. Environ. Sci. Technol. 2015, 49, (13), 8078-8086. 22. Collin, B.; Tsyusko, O. V.; Starnes, D. L.; Unrine, J. M., Effect of natural organic matter on dissolution and toxicity of sulfidized silver nanoparticles to Caenorhabditis elegans. Environ. Sci. Nano 2016, 3, (4), 728-736. 23. Yang, X. Y.; Jiang, C. J.; Hsu-Kim, H.; Badireddy, A. R.; Dykstra, M.; Wiesner, M.; Hinton, D. E.; Meyer, J. N., Silver nanoparticle behavior, uptake, and toxicity in Caenorhabditis elegans: Effects of natural organic matter. Environ. Sci. Technol. 2014, 48, (6), 3486-3495. 24. Elemental Compositions and Stable Isotopic Ratios of IHSS Samples. http://humicsubstances.org/elemental-compositions-and-stable-isotopic-ratios-of-ihss-samples/ (October 09). 25. Mousavi, M. P. S.; Gunsolus, I. L.; De Jesus, C. E. P.; Lancaster, M.; Hussein, K.; Haynes, C. L.; Buhlmann, P., Dynamic silver speciation as studied with fluorous-phase ionselective electrodes: Effect of natural organic matter on the toxicity and speciation of silver. Sci. Total Environ. 2015, 537, 453-461. 26. Battocchio, C.; Meneghini, C.; Fratoddi, I.; Venditti, I.; Russo, M. V.; Aquilanti, G.; Maurizio, C.; Bondino, F.; Matassa, R.; Rossi, M.; Mobilio, S.; Polzonetti, G., Silver


29


620 621 622 623 624 625 626 627 628 629 630 631 632 633 634 635 636 637 638 639 640 641 642 643 644 645 646 647 648 649 650 651 652 653 654 655 656 657 658 659 660 661 662 663 664

Page 30 of 32

nanoparticles stabilized with thiols: A close look at the local chemistry and chemical structure. J. Phys. Chem. C 2012, 116, (36), 19571-19578. 27. Adams, N. W. H.; Kramer, J. R., Reactivity of Ag+ ion with thiol ligands in the presence of iron sulfide. Environ. Toxicol. Chem. 1998, 17, (4), 625-629. 28. Metreveli, G.; Frombold, B.; Seitz, F.; Grun, A.; Philippe, A.; Rosenfeldt, R. R.; Bundschuh, M.; Schulz, R.; Manz, W.; Schaumann, G. E., Impact of chemical composition of ecotoxicological test media on the stability and aggregation status of silver nanoparticles. Environ. Sci. Nano 2016, 3, (2), 418-433. 29. Romer, I.; Gavin, A. J.; White, T. A.; Merrifield, R. C.; Chipman, J. K.; Viant, M. R.; Lead, J. R., The critical importance of defined media conditions in Daphnia magna nanotoxicity studies. Toxicol. Lett. 2013, 223, (1), 103-108. 30. Smith, K. S.; Balistrieri, L. S.; Todd, A. S., Using biotic ligand models to predict metal toxicity in mineralized systems. Appl. Geochem. 2015, 57, 55-72. 31. Mulholland, P. J., Dissolved organic matter concentration and flux in streams. J. North Am. Benthological Soc. 1997, 16, 131-141. 32. Topuz, E.; Sigg, L.; Talinli, I., A systematic evaluation of agglomeration of Ag and TiO2 nanoparticles under freshwater relevant conditions. Environ. Pollut. 2014, 193, 37-44. 33. OECD Guideline for testing of chemicals, Test No. 202: Daphnia sp., acute immobilisation test. Organization of Economic Co-operation and Development (OECD). 34. Gottschalk, F.; Sun, T. Y.; Nowack, B., Environmental concentrations of engineered nanomaterials: Review of modeling and analytical studies. Environ. Pollut. 2013, 181, 287-300. 35. Biological test method: Reference method for determining acute lethality of effluents to Daphnia magna. Environment Canada Environmental Protection Publications, Ottawa 1990. 36. Vonzelewsky, A.; Barbosa, L.; Schlapfer, C. W., Poly(ethylenimines) as Brønsted bases and as ligands for metal ions. Coord. Chem. Rev. 1993, 123, (1-2), 229-246. 37. Kawaguchi, T.; Hasegawa, M., Structure of dextran–magnetite complex: Relation between conformation of dextran chains covering core and its molecular weight. J. Mater. Sci. Mater. Med. 2000, 11, (1), 31-35. 38. Regmi, R.; Gumber, V.; Rao, V. S.; Kohli, I.; Black, C.; Sudakar, C.; Vaishnava, P.; Naik, V.; Naik, R.; Mukhopadhyay, A., Discrepancy between different estimates of the hydrodynamic diameter of polymer-coated iron oxide nanoparticles in solution. J. Nanopart. Res. 2011, 13, (12), 6869-6875. 39. Bihari, P.; Vippola, M.; Schultes, S.; Praetner, M.; Khandoga, A. G.; Reichel, C. A.; Coester, C.; Tuomi, T.; Rehberg, M.; Krombach, F., Optimized dispersion of nanoparticles for biological in vitro and in vivo studies. Part. fibre. toxicol. 2008, 5, (1), 1. 40. Murdock, R. C.; Braydich-Stolle, L.; Schrand, A. M.; Schlager, J. J.; Hussain, S. M., Characterization of nanomaterial dispersion in solution prior to in vitro exposure using dynamic light scattering technique. Toxicol. Sci. 2008, 101, (2), 239-253. 41. Palomino, D.; Stoll, S., Fulvic acids concentration and pH influence on the stability of hematite nanoparticles in aquatic systems. J. Nanopart. Res. 2013, 15, (2). 42. Tan, S.; Erol, M.; Attygalle, A.; Du, H.; Sukhishvili, S., Synthesis of positively charged silver nanoparticles via photoreduction of AgNO3 in branched polyethyleneimine/HEPES solutions. Langmuir 2007, 23, (19), 9836-9843. 43. Leenheer, J. A.; Croué, J.-P., Characterizing aquatic dissolved organic matter. Environ. Sci. Technol. 2003, 37, (1), 18A-26A.


30

Page 31 of 32

665 666 667 668 669 670 671 672 673 674 675 676 677 678 679 680 681 682 683 684 685 686 687 688 689 690 691 692 693 694 695 696 697 698 699 700 701 702 703 704 705 706


44. Piccolo, A., The supramolecular structure of humic substances. Soil science 2001, 166, (11), 810-832. 45. Adegboyega, N. F.; Sharma, V. K.; Siskova, K.; Zboril, R.; Sohn, M.; Schultz, B. J.; Banerjee, S., Interactions of aqueous Ag+ with fulvic acids: Mechanisms of silver nanoparticle formation and investigation of stability. Environ. Sci. Technol. 2013, 47, (2), 757-764. 46. Noguez, C., Surface plasmons on metal nanoparticles: The ifluence of shape and physical environment. J. Phys. Chem. C 2007, 111, (10), 3806-3819. 47. Sikora, F. J.; Stevenson, F. J., Silver complexation by humic substances conditional stability constants and nature of reactive sites. Geoderma 1988, 42, (3-4), 353-363. 48. Liu, J. Y.; Sonshine, D. A.; Shervani, S.; Hurt, R. H., Controlled release of biologically active silver from nanosilver surfaces. ACS Nano 2010, 4, (11), 6903-6913. 49. Liu, J. Y.; Hurt, R. H., Ion release kinetics and particle persistence in aqueous nano-silver colloids. Environ. Sci. Technol. 2010, 44, (6), 2169-2175. 50. Li, X.; Lenhart, J. J.; Walker, H. W., Aggregation kinetics and dissolution of coated silver nanoparticles. Langmuir 2012, 28, (2), 1095-1104. 51. Kittler, S.; Greulich, C.; Diendorf, J.; Koller, M.; Epple, M., Toxicity of silver nanoparticles increases during storage because of slow dissolution under release of silver ions. Chem. Mat. 2010, 22, (16), 4548-4554. 52. Dobias, J.; Bernier-Latmani, R., Silver release from silver nanoparticles in natural waters. Environ. Sci. Technol. 2013, 47, (9), 4140-4146. 53. Zhang, H. Y.; Smith, J. A.; Oyanedel-Craver, V., The effect of natural water conditions on the anti-bacterial performance and stability of silver nanoparticles capped with different polymers. Water Res. 2012, 46, (3), 691-699. 54. Nasser, F.; Lynch, I., Secreted protein eco-corona mediates uptake and impacts of polystyrene nanoparticles on Daphnia magna. J. proteomics 2016, 137, 45-51. 55. Lee, Y. J.; Kim, J.; Oh, J.; Bae, S.; Lee, S.; Hong, I. S.; Kim, S. H., Ion-release kinetics and ecotoxicity effects of silver nanoparticles. Environ. Toxicol. Chem. 2012, 31, (1), 155-159. 56. Blinova, I.; Niskanen, J.; Kajankari, P.; Kanarbik, L.; Käkinen, A.; Tenhu, H.; Penttinen, O.-P.; Kahru, A., Toxicity of two types of silver nanoparticles to aquatic crustaceans Daphnia magna and Thamnocephalus platyurus. Environ. Sci. Pollut. Res. 2013, 20, (5), 3456-3463. 57. Allen, H. J.; Impellitteri, C. A.; Macke, D. A.; Heckman, J. L.; Poynton, H. C.; Lazorchak, J. M.; Govindaswamy, S.; Roose, D. L.; Nadagouda, M. N., Effects from filtration, capping agents, and presence/absence of food on the toxicity of silver nanoparticles to Daphnia magna. Environ. Toxicol. Chem. 2010, 29, (12), 2742-2750. 58. Jung, Y.; Park, C.-B.; Kim, Y.; Kim, S.; Pflugmacher, S.; Baik, S., Application of multispecies microbial bioassay to assess the effects of engineered nanoparticles in the aquatic environment: Potential of a luminous microbial array for toxicity risk assessment (LumiMARA) on testing for surface-coated silver nanoparticles. Int. J. Environ. Res. Public Health 2015, 12, (7), 8172-8186. 59. Bozich, J. S.; Lohse, S. E.; Torelli, M. D.; Murphy, C. J.; Hamers, R. J.; Klaper, R. D., Surface chemistry, charge and ligand type impact the toxicity of gold nanoparticles to Daphnia magna. Environ. Sci. Nano 2014, 1, (3), 260-270.

707 708

For Table of Contents Only


31


Page 32 of 32

709


32

Implications of Pony Lake Fulvic Acid for the Aggregation and

Recommend Documents