Relative Contributions of Copper Oxide ... - ACS Publications

Subscriber access provided by University of Newcastle, Australia

Article

Relative Contributions of Copper Oxide Nanoparticles and Dissolved Copper to Cu Uptake Kinetics of Gulf Killifish (Fundulus grandis) Embryos Chuanjia Jiang, Benjamin T. Castellon, Cole W. Matson, George R. Aiken, and Heileen Hsu-Kim Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.6b04672 • Publication Date (Web): 12 Jan 2017 Downloaded from http://pubs.acs.org on January 15, 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 30

Environmental Science & Technology

Relative Contributions of Copper Oxide Nanoparticles and Dissolved Copper to Cu Uptake Kinetics of Gulf Killifish (Fundulus grandis) Embryos Chuanjia Jiang 1,2†, Benjamin T. Castellon 2,3, Cole W. Matson 2,3, George R. Aiken 4, Heileen Hsu-Kim 1,2*

1

Department of Civil and Environmental Engineering, Duke University, Durham, NC 27708

2

Center for the Environmental Implications of NanoTechnology (CEINT), Duke University,

Durham, NC 27708 3

Department of Environmental Science, Institute of Biomedical Studies, Center for Reservoir

and Aquatic Systems Research (CRASR), Baylor University, Waco, TX 76798 4

U.S. Geological Survey, Boulder, CO 80303

*Corresponding author: Heileen Hsu-Kim. Email: [email protected]. Phone +1-919-660-5109

†

Present address: State Key Laboratory of Advanced Technology for Materials Synthesis and

Processing, Wuhan University of Technology, Luoshi Road 122, Wuhan 430070, China

1 ACS Paragon Plus Environment


1

Abstract: The toxicity of soluble metal-based nanomaterials may be due to the uptake of metals

2

in both dissolved and nanoparticulate forms, but the relative contributions of these different

3

forms to overall metal uptake rates under environmental conditions are not quantitatively defined.

4

Here we investigated the linkage between the dissolution rates of copper(II) oxide (CuO)

5

nanoparticles (NPs) and their bioavailability to Gulf killifish (Fundulus grandis) embryos, with

6

the aim of quantitatively delineating the relative contributions of nanoparticulate and dissolved

7

species for Cu uptake. Gulf killifish embryos were exposed to dissolved Cu and CuO NP

8

mixtures comprising a range of pH values (6.3–7.5) and three types of natural organic matter

9

(NOM) isolates at various concentrations (0.1–10 mg-C L−1), resulting in a wide range of CuO

10

NP dissolution rates that subsequently influenced Cu uptake. First-order dissolution rate

11

constants of CuO NPs increased with increasing NOM concentration and for NOM isolates with

12

higher aromaticity, as indicated by specific ultraviolet absorbance (SUVA), while Cu uptake rate

13

constants of both dissolved Cu and CuO NP decreased with NOM concentration and aromaticity.

14

As a result, the relative contribution of dissolved Cu and nanoparticulate CuO species for the

15

overall Cu uptake rate was insensitive to NOM type or concentration but largely determined by

16

the percentage of CuO that dissolved. These findings highlight SUVA and aromaticity as key

17

NOM properties affecting the dissolution kinetics and bioavailability of soluble metal-based

18

nanomaterials in organic-rich waters. These properties could be used in the incorporation of

19

dissolution kinetics into predictive models for environmental risks of nanomaterials.


Page 2 of 30

Page 3 of 30

20


Graphic for the Table of Contents

21

22



23

INTRODUCTION

24

With the rapid development of nanotechnology in the last few decades, there have been

25

increasing concerns over the implications of engineered nanomaterials to human health and the

26

environment.1-6 To elicit toxicity in an organism, nanomaterials and associated toxic species

27

must first be taken up by the organism or approach its immediate vicinity.7 Therefore,

28

quantifying the bioavailability of nanomaterials and constituents that may be released from

29

nanomaterials is a key step in explaining the toxic effects of nanomaterials in various

30

environmental and physiological media.

31

After release into the environment, nanomaterials can undergo a range of processes that

32

alter their bioavailability and toxicity. These processes include homoaggregation and

33

heteroaggregation, dissolution and release of ionic species, surface modifications, and

34

transformation into materials with a different chemical composition (e.g., sulfidation).8-14 The

35

relative rates of these concurrent processes, and not necessarily their reaction potential at

36

thermodynamic equilibrium, will control mobilization of nanomaterials in the aquatic

37

environment and exposure to target organisms. For soluble metal-based nanomaterials such as

38

silver (Ag), zinc oxide (ZnO) and copper oxide (CuO) nanoparticles (NPs), both nanoparticulate

39

and dissolved forms of the metals can be taken up by the organisms and typically at different

40

rates.15-19 Moreover, the dissolution and bioavailability of metal-based nanomaterials can be

41

strongly influenced by the physico-chemical properties of the media, notably pH,11, 20 ionic

42

strength/salinity,21, 22 hardness,23 and inorganic and organic ligands (e.g. chloride, amino acids,

43

extracellular polymeric substances and humic substances).11, 12, 24-30 Due to these influencing

44

factors as well as differences in nanomaterial properties and test organisms, there have been

45

apparently inconsistent reports where, in some studies, the undissolved nanomaterials are


Page 4 of 30

Page 5 of 30


46

bioavailable and responsible for the observed toxicity,31-33 while in other cases, the toxic effects

47

could be attributed solely to the uptake of dissolved metal ions.34-36 These apparent

48

inconsistencies highlight a major gap in the field of nanoecotoxicology, in which the relative

49

contributions of nanoparticles and dissolved ions released from the NPs have not been

50

effectively determined. In this respect, the relative uptake rates of NPs and corresponding

51

dissolved metal ions need to be quantified under environmentally relevant conditions.

52

While the bioavailability of metal ions has been extensively studied,37, 38 efforts have

53

recently been made to quantify uptake kinetics of NPs by experimentation (e.g. radioactive and

54

stable isotope labeling)22, 39 and modeling (e.g. biodynamic models).15, 17, 22, 40 To increase the

55

accuracy of measured biodynamic parameters of soluble metal-based NPs, it is desirable to

56

account for the dissolution of the NPs prior to41 or during uptake experiments,22, 39 using

57

measured or assumed dissolved metal concentrations.22, 39 Moreover, other constituents of the

58

aquatic environment such as dissolved natural organic matter (NOM) are known to influence

59

nanoparticle surface chemistry and bioavailability in complex ways, e.g. through adsorption of

60

NOM to particle surfaces, change in colloidal aggregation rates, enhanced dissolution of

61

nanoparticles, and metal-ligand complexation of metal ions released from the nanoparticles.42

62

These processes have been extensively studied, but rarely in combination and generally without

63

quantification of reaction rates that are needed to delineate the relative importance of

64

nanoparticles and dissolved metals for organismal exposures.

65

This study aimed to quantify the relative rates of nanoparticle and dissolved metal uptake

66

and the importance of NOM type and concentration for controlling this balance. Herein, we

67

evaluated the copper uptake kinetics in Gulf killifish (Fundulus grandis) embryos exposed to

68

CuO NPs in diluted artificial seawater (ASW). CuO NPs were selected as a model soluble metal-



69

based nanomaterial, for their moderate solubility, which allows for relatively low exposure doses,

70

as well as the wide application of CuO NPs and other Cu-based nanomaterials in biocidal agents

71

and coatings.11, 12, 43, 44 While zebrafish and their embryos have been widely used as model

72

organisms in nanotoxicological studies,12, 18, 25, 34 Gulf killifish and its closely related sister

73

species Atlantic killifish (Fundulus heteroclitus) are widely distributed in estuarine habitats on

74

the eastern coasts of the United States. Because of their wide ecological distribution as well as

75

their tolerance of salinity and other environmental gradients, these species are commonly used as

76

model teleosts in ecotoxicological and, most recently, nanotoxicological studies.21, 45, 46

77

MATERIALS AND METHODS

78

Overall Experimental Design. The addition of CuO NPs into the test media results in a

79

mixture of NPs and dissolved Cu, the concentration ratio of which changes with time. Thus, in

80

addition to CuO NP uptake experiments, separate positive control experiments with dissolved Cu

81

were performed in order to quantify uptake rates of dissolved Cu. Copper exposure experiments

82

were performed with three types of NOM and a range of NOM concentrations and pH values

83

(Table 1) that were selected to produce a wide spectrum of nanoparticle aggregation and

84

dissolution rates. Dissolution kinetics of CuO NPs were determined at high time resolution using

85

anodic stripping voltammetry (ASV),47, 48 under identical conditions as the uptake experiments.

86

Collectively these data were used to differentiate uptake of dissolved Cu and nanoparticulate

87

CuO in the CuO NP exposures, in order to determine the relative contributions of dissolved and

88

nanoparticulate species to Cu uptake kinetics across a wide range of water chemistry conditions.

89

Materials and Test Organism. Purity and sources of chemicals are provided in Table S1

90

in Supporting Information (SI). CuO NPs (nominally 25 min prior to use, resulting in Z-average hydrodynamic diameter

106

of 250±30 nm via dynamic light scattering (DLS) (Malvern Zetasizer Nano ZS).

107

Three aquatic NOM isolates were used in the study, including Suwannee River humic

108

and fulvic acids (SRHA, 2S101H; SRFA, 2S101F) from the International Humic Substances

109

Society and Williams Lake hydrophobic acid (WLHpoA). These three NOM isolates were

110

selected because our previous work has shown a wide range of NOM-enhanced dissolution rates

111

for ZnO NPs.48 These NOM isolates also span the typical range of specific UV absorbance at

112

280 nm (SUVA280), a parameter that is positively correlated with dissolution rates:

113

4.92±0.34 L mg-C−1 m−1 for SRHA, 3.34±0.22 L mg-C−1 m−1 for SRFA and 1.22±0.11 L mg-C−1

114

m−1 for WLHpoA.48 Dried powders of the NOM isolates were dissolved in ultrapure water, and



115

the pH was adjusted to 6.5–7.5 with 0.1 N NaOH. The stocks were then filtered through 0.2 µm

116

nylon-membrane syringe filters (VWR), and the concentration of the NOM stocks (typically 170

117

to 400 mg-C L−1) was determined using a Total Organic Carbon (TOC) Analyzer (TOC-L CPN,

118

Shimadzu).

119

Embryos used in the study were produced by wild-type F. grandis populations collected

120

from two reference sites on Galveston Bay, Texas: Smith Point (29°32'37.26"N, 94°47'08.12"W)

121

and Gangs Bayou (29°15'30.34"N, 94°54'45.00"W). These populations were established and

122

maintained as previously described.49 Briefly, the killifish were kept in recirculating 10‰ ASW

123

(Instant Ocean) on a 14:10 h light cycle at 22–25 ºC and were fed twice daily (Purina AquaMax).

124

After spawning, the fertilized embryos were disinfected with 0.3% hydrogen peroxide in 5‰

125

ASW, rinsed twice, and incubated at 28°C in 5‰ ASW prior to use in uptake experiments at

126

48±2 h post-fertilization (hpf). The embryos were on average 2.0–2.1 mm in diameter and 6 mg

127

wet weight. Average dry weight of the embryos at 48 hpf was 0.85±0.04 mg, as measured after

128

drying at 60 ºC overnight (> 12 h). The maintenance and handling of Gulf killifish were

129

performed in compliance with the relevant laws and institutional guidelines and approved by the

130

Baylor University Institutional Animal Care and Use Committee.

131

CuO NP and Dissolved Copper Uptake Experiments. All uptake experiments were

132

performed in dilute ASW (5‰ salinity) buffered to pH 6.3–7.5 with 2 mM MOPS and filtered

133

through a 0.2-µm polycarbonate membrane. Experiments were performed with or without an

134

NOM isolate at 0.1–10 mg-C L−1. In each experiment, buffered ASW and NOM stocks were

135

pipetted into five 60-mL flat-bottom polypropylene tubes (SC480-W, Environmental Express),

136

and then 4 viable embryos at 48±2 hpf were placed into each tube. Subsequently, CuO NP stock

137

or dissolved Cu(NO3)2 solution (1.0 or 0.1 mM) was added, achieving a total volume of 40 mL in


Page 8 of 30

Page 9 of 30


138

each tube. The suspensions were mixed end over end twice and incubated at 27–28 ºC in the dark

139

for up to 48 h (IsotempTM, Model 6841, Fisher Scientific). Total nominal Cu concentration in the

140

samples was 30 µM for CuO NP exposure and 1.0 or 0.4 µM (for experiments at pH 7.5) for

141

dissolved Cu exposure. A blank control sample was prepared in the same way as described above

142

except that no Cu was added.

143

Each copper test mixture (dissolved Cu or CuO NP; with or without NOM) was

144

performed in duplicate, with a few exceptions in triplicate. Copper content in the embryos was

145

measured at different times post dosing (ranging from 2 h to 48 h). At each time point, aliquots

146

of the exposure media were taken from above the embryos for measurements of total Cu

147

concentration in the media (i.e. [Cu]T) by ICP-MS and, for CuO NP exposure, hydrodynamic

148

diameter (dh) and zeta potential (ζ-potential) of the CuO NPs (Malvern Zetasizer Nano ZS). Then

149

the 4 embryos in the sample were removed and rinsed in 5 mL of 5‰ ASW for 2 min,

150

transferred to a clean 60-mL polypropylene tube (SC480-W, Environmental Express) and stored

151

at −20 ºC. In some experiments, the embryos were dechorionated mechanically after exposure,21

152

and the removed chorions (i.e., the acellular outer membrane enveloping the embryo) were

153

collected in a 60-mL polypropylene tube. At the 48-h time point (i.e. 96±2 hpf), the embryos

154

during rinsing were examined under a light microscope (AmScope, SE305R-PZ) to ensure that

155

the embryos were still viable, as indicated by cardiac rhythm. The embryos or removed chorions were digested in 2 mL of concentrated nitric acid (~70%

156 157

as HNO3) at 90 ºC for 3 h on a hot block (Environmental Express, SC100), prior to the

158

determination of Cu content in the embryos/chorions ([Cu]org, in unit of molCu per embryo;

159

justification given in the SI) by ICP-MS. More details of the uptake experiment are provided in

160

the SI.



161

Page 10 of 30

CuO NP Dissolution Experiments. The dissolution kinetics of CuO NPs were

162

monitored over 48 to 72 h in duplicate or triplicate experiments, under the same conditions as the

163

uptake experiments, but typically with no embryos in the samples. In each experiment, 10 or 11

164

CuO NP samples (with 30 µM nominal total Cu) were prepared in 60-mL flat-bottom

165

polypropylene tubes using the same method as the CuO NP uptake experiments, and total

166

dissolved Cu concentration (i.e. [Cu]d) was measured by ASV at multiple time points (from 1 to

167

72 h). A few experiments were performed with embryos to test if they affected the dissolution

168

kinetics of CuO NPs. In addition to the water chemistry conditions in Table 1, CuO NP

169

dissolution was also measured at additional pH values of 6.5, 6.8 and 7.2, with 1 mg-C L−1

170

SRHA. The ASV measurement was performed on a Metrohm 663 VA Stand and Ecochemie

171

µAutolab Type III potentiostat system,47, 48 with detailed procedures provided in the SI.

172

Model of CuO NP Dissolution Kinetics. Dissolution kinetics of CuO NPs were modeled

173

as surface-controlled dissolution, and the change in [Cu]d over time (t) was described using

174

Equation 1:

175

 d [Cu]d [Cu]d  = k d ([CuO]0 − [Cu]d )1 −  [Cu]  dt d,eq  

(1)

176

where kd is the apparent first-order dissolution rate constant (in molCu molCuO−1 s−1) of CuO NP,

177

[CuO]0 is the initial total concentration of CuO NP (i.e., 30 µM), and [Cu]d,eq is the dissolved Cu

178

concentration measured at equilibrium (defined at t = 48 or 72 h). The rate constant kd was

179

estimated by fitting Equation 1 to the measured dissolved Cu data for time points between 1 and

180

24.2 h, where [Cu]d had not reached dissolution equilibrium. All data fitting was performed with

181

Berkeley Madonna software v.8.3.18), using the Runge-Kutta (RK4) method with a step size of

182

0.02 h. Initial dissolved Cu concentration in the reaction mixture (0.3–1.7 µM) was designated

183

by the [Cu]d measured at t = 1 h or the average of [Cu]d measured at 1 and 2 h, if the former was 10 ACS Paragon Plus Environment

Page 11 of 30


184

higher than the latter. For dissolution experiments performed at pH 7.5 or with 0.1 mg-C L−1

185

WLHpoA, the first 12 h of the dissolution curves were linearly fitted instead of using Equation 1

186

(reasons given below in the subsection CuO NP Dissolution Rates).

187

Model of Dissolved Cu and CuO NP Uptake Kinetics. The change in Cu content in

188

Gulf killifish embryos ([Cu]org, in molCu per embryo) exposed to dissolved Cu and CuO NP over

189

time was due to uptake/excretion of dissolved Cu ions and/or undissolved CuO NPs. This

190

process was described by Equation 2 for organisms exposed to dissolved Cu and Equation 3 for

191

organisms exposed to CuO NPs (which partly dissolved in the test matrices):

192

d [Cu]org dt

193

d [Cu]org dt

= k u,d [Cu]T − k e [Cu]org

(2)

= k u,NP ([Cu]T − [Cu]d ) + k u,d [Cu]d − k e' [Cu]org

(3)

194

where ku,d and ku,NP are the uptake rate constants (L embryo−1 h−1) of dissolved Cu and CuO NPs,

195

respectively, while ke and k’e are the elimination rate constants (h−1) of Cu in embryos exposed to

196

dissolved Cu and CuO NPs, respectively. Herein, we did not distinguish between the elimination

197

rates of dissolved Cu and nanoparticulate CuO, since their respective concentrations in the

198

embryos could not be measured. Neither did we account for the possibility of biphasic metal

199

elimination, which features fast and slow elimination phases.17

200

Thus, the model was fit to the first 24 or 12 h of experimental data, when [Cu]org was

201

relatively low and elimination of Cu from the embryos was assumed to be negligible. This

202

assumption resulted in modifications of Equations 2 and 3 to the following:

203

d [Cu]org dt

204

t WLHpoA, as shown in Table 1) were more effective at

231

promoting the dissolution of the CuO NPs, which is consistent with previous findings on NOM-

232

promoted dissolution of bulk CuO50 as well as ZnO NPs.48 At a low concentration of 0.1 mg-C

233

L−1, WLHpoA inhibited CuO NP dissolution within the first 12 h, which is consistent with

234

decreased dissolution of ZnO NP (6.14 mM total concentration) by relatively low concentration

235

(2.6 mg-C L−1) of SRFA.51

236

It should be noted that all dissolution data were obtained by ASV. This approach for

237

dissolved Cu ([Cu]d) measurement was validated in a subset of mixtures by ultracentrifugation of

238

the CuO NP suspensions and quantification of Cu in the supernatant by ICP-MS. For example,

239

[Cu]d values measured by ASV was 109% and 88% of the ICP-MS values, for CuO NP

240

suspensions in buffered 5‰ ASW with no NOM and with 5 mg-C L−1 SRFA, respectively

241

(Figure S3). The presence of 4 embryos in the CuO NP suspensions did not appear to affect the

242

dissolution kinetics of CuO NPs, as shown in controls with buffered 5‰ ASW with and without

243

1 mg-C L−1 SRHA (Figure S4) (These are conditions that could be expected to be most

244

susceptible to dissolution changes caused by the embryos). It should also be noted that the

245

starting dissolved Cu concentration (at time 0) could not be measured, due to protocol limitations.

246

Fast initial dissolution was inferred from the measurements at 1 h, where the first measured [Cu]d

247

could be a quarter of maximum [Cu]d under the same conditions (e.g. with 10 mg-C −1 SRHA).

248

Such patterns, that is, faster initial dissolution followed by slower dissolution was also observed

249

for CuO NPs in DI water or mesocosm freshwater, as measured by membrane dialysis.52



250 251 252

Table 1. CuO NP dissolution rate constants and uptake rate constants for F. grandis embryos exposed to mixtures of dissolved Cu and CuO NPs under a variety of water chemistry conditions that included three types of natural organic matter (NOM) isolates with a spectrum of values for specific UV absorbance at 280 nm (SUVA280). Uptake rate Apparent CuO Uptake rate constant of dissolution rate constant of nanoparticulate NOM type and CuO constant dissolved copper No. pH (range) concentration [DOC]×SUVA280 kd ku,d ku,NP −1 −1 −1 −1 −1 (m ) (molCu molCuO s ) (L embryo h ) (L embryo−1 h−1) 1 7.0 (7.0–7.1) no NOM 0 0.71×10−6 64.6 (±3.7) ×10−6 3.9 (±1.5) ×10−6 2 7.0 (7.0–7.1) 0.1 mg-C L−1 WLHpoA 0.12 n/a 71.1 (±29.5) ×10−6 4.8 (±0.4) ×10−6 3 7.0 (7.0–7.1) 5 mg-C L−1 WLHpoA 6.1 0.65×10−6 43.2 (±4.6) ×10−6 2.0 (±0.1) ×10−6 4 6.3 (6.3–6.4) 1 mg-C L−1 SRHA 4.9 2.5×10−6 36.3 (±6.9) ×10−6 3.0 (±0.4) ×10−6 5 7.0 (7.0–7.1) 1 mg-C L−1 SRHA 4.9 0.56×10−6 30.1 (±2.1) ×10−6 1.4 (±0.5) ×10−6 6 7.5 (7.4–7.5) 1 mg-C L−1 SRHA 4.9 n/a 25.8 (±8.1) ×10−6 1.3 (±0.2) ×10−6 7 7.0 (7.0–7.1) 5 mg-C L−1 SRFA 16.7 1.1×10−6 24.3 (±4.0) ×10−6 1.2 (±0.3) ×10−6 8 7.0 (7.0–7.1) 5 mg-C L−1 SRHA 24.6 2.6×10−6 17.7 (±1.7) ×10−6 1.4 (±0.9) ×10−6 9 7.0 (7.0–7.1) 10 mg-C L−1 SRHA 49.2 2.2×10−6 11.2 (±0.6) ×10−6 0.8 (±0.3) ×10−6


Page 14 of 30

Page 15 of 30


(a)

Dissolved Cu [Cu]d (µM)

5

4

1 mg-C L-1 SRHA pH 6.3, kd=2.5×10-6

3

pH 6.5, kd=1.8×10-6 pH 6.8, kd=1.1×10-6 pH 7.0, kd=5.6×10-7

2

pH 7.2, kd=5.7×10-7 pH 7.5, kd: n/a

1

-1

Unit of kd: molCu molCuO

s

-1

0 0

10

20

30

40

50

60

70

Time (h)

253

Dissolved Cu [Cu]d (µM)

5

(b)

4 10 mg-C L-1 SRHA 5 mg-C L-1 SRHA 5 mg-C L-1 SRFA 1 mg-C L-1 SRHA no NOM 5 mg-C L-1 WLHpoA 0.1 mg-C L-1 WLHpoA

3

2

1

pH = 7.0

0 0

254 255 256 257 258 259

10

20

30

40

50

60

70

Time (h)

Figure 1. Measured (symbols) and modeled (lines) dissolution of CuO NPs in 5‰ ASW (a) buffered to pH 6.3–7.5 in the presence of 1 mg-C L−1 SRHA, and (b) buffered at pH 7.0 in the presence or absence of different NOM isolates (SRHA, SRFA and WLHpoA) at 0.1–10 mg-C L−1. Ionic strength ~90 mM, T = 28 ºC. Experiments were performed in the dark. Error bars are standard deviation of replicate experiments (N = 2 or 3).



260

Colloidal Stability of CuO NPs. In 5‰ ASW (~90 mM ionic strength) with no NOM or

261

a very low NOM concentration (e.g. 0.1 mg-C L−1 WLHpoA), the CuO NPs quickly aggregated,

262

with Z-average hydrodynamic diameter (dh) reaching 2 µm within 12 h (Figure 2a). The addition

263

of higher SUVA280 NOM (e.g. SRHA at 1 mg-C L−1 and greater) or at higher concentration of

264

low SUVA280 NOM (e.g. 5 mg-C L−1 WLHpoA) stabilized the NPs against aggregation (Figure

265

2a and complete data set in Figure S5). Expectedly, the larger NP aggregates settled at faster

266

rates, with less total copper remaining in the water column (Figures 2b and S6). For example,

267

after 24 h of incubation, 16±2 and 23±1 µM Cu remained in media without or with 1 mg-C L−1

268

SRHA, respectively, as measured at 9 mm above the bottom of the containers (Figure 2b). When

269

SRHA concentration increased from 1 to 10 mg-C L−1, the aggregation and sedimentation

270

behaviors of the CuO NPs were not further affected. The increased colloidal stability of the CuO

271

NPs was likely due to the steric hindrance of aggregation by the NOM,48, 53, 54 and, to a lesser

272

extent, the more negative surface charge. The surface charge of the CuO NPs became more

273

negative as NOM concentration or SUVA280 increased (Figures 2c and S7). With no NOM, the ζ-

274

potential of the CuO NPs was on average −5±1 mV in the 1 h to 24 h timeframe. With the

275

addition of 1 and 10 mg-C L−1 SRHA, the ζ-potential was reduced to −10±1 and −14±1 mV,

276

respectively. Considering the different dissolution kinetics and colloidal stabilities of CuO NPs,

277

we expected the uptake kinetics of CuO NPs to vary between the tested water chemistry

278

conditions.


Page 16 of 30

Page 17 of 30


2000 1500

dh (nm)

1000 500 300 200 100 0

280 281 282 283 284

0

10

20

30

Time (h)

40

50

0

(b)

30 25

(c)

-4 ζ−potential (mV)

no NOM 0.1 mg-C L-1 WLHpoA 5 mg-C L-1 WLHpoA 5 mg-C L-1 SRFA 1 mg-C L-1 SRHA 5 mg-C L-1 SRHA 10 mg-C L-1 SRHA

(a)

2500

Total Cu concentration in media [Cu]T (µM)

279

20 15 10

-8

-12 5 0

0

10

20

30

40

Time (h)

50

-16

0

10

20

30

40

50

Time (h)

Figure 2. (a) Z-average hydrodynamic diameter (dh), (b) total suspended Cu concentration ([Cu]T) and (c) zeta potential (ζ-potential) of the CuO NPs in 5‰ ASW buffered to pH 7.0 with 2 mM MOPS, in the absence or presence of different NOM isolates at different concentrations. Ionic strength ~90 mM, T = 28 ºC. Experiments were performed in the dark. Error bars are standard deviations of replicate CuO NP uptake experiments (N = 2 to 6).



285

Uptake and Distribution of Dissolved Cu and CuO NPs. The copper content in Gulf

286

killifish embryos increased with time after exposure to dissolved Cu (Figure S8) or CuO NPs

287

(Figure S9). The ratio of Cu content in chorions to that measured in whole embryos (i.e., the

288

chorion and the contents enclosed within) was typically > 0.4 (Figure S10), suggesting that a

289

major portion of the Cu was associated with the chorions. This observation is consistent with

290

silver uptake in Atlantic killifish and zebrafish embryos exposed to Ag NPs or AgNO3, where the

291

majority (> 0.6) of measured Ag in exposed embryos was also associated with chorions.18, 21

292

The Cu distribution in the embryos was time-dependent. For both dissolved Cu and CuO

293

NP exposures, Cu content measured in whole embryos typically increased throughout the 48-h

294

dosing period, while Cu content in chorions increased at slower rates after 12 or 24 h post dosing

295

(Figures S8 and S9). Consequently, the ratio of Cu content in chorions to that measured in whole

296

embryos typically decreased after 12 or 24 h (Figure S10), indicating that Cu became saturated in

297

the chorions over time and more passed through into the interior compartments.

298

We note that at some time points the calculated ratios were >1, which was due to

299

biological variability and measurement uncertainty. We also recognize that Cu accumulation in

300

the chorion may not be directly accessible for specific organism functions or will induce toxic

301

effects. In fact, the chorion can often serve as a barrier for nanoparticles.18 However, uptake into

302

the chorion is a first step towards bioaccumulation, and the chorion loaded with nanoparticles

303

can become a site for long-term metal exposure and in situ biotransformations.21 Thus, the

304

differentiation of dissolved and nanoparticulate Cu uptake pathways provide useful insights to

305

understanding the bioavailability of metals originating from nanomaterials.

306 307

Uptake Kinetics of Dissolved Cu and CuO NPs. The modeled Cu uptake kinetics generally agreed with measured data within the first 24 or 12 h post dosing (Figures S8 and S9).


Page 18 of 30

Page 19 of 30


308

In modeling the uptake kinetics (Equations 4 and 5), both input parameters (as continuous

309

functions of time) of dissolved Cu concentration ([Cu]d) and total Cu concentration in suspension

310

([Cu]T) changed with time. The decrease in [Cu]T for CuO NP uptake experiments was mainly

311

due to the sedimentation of the NPs (Figures 2b and S6), while the change for the dissolved Cu

312

uptake experiments (Figure S11) was due to Cu uptake into embryos. [Cu]d versus time was

313

typically modeled according to Equation 1. The [Cu]T values in both CuO NP and dissolved Cu

314

uptake experiments were assumed to change linearly with time within the first 24 h, and the

315

linearly fitted equations of [Cu]T versus time were used as input parameters in modeling the

316

uptake kinetics.

317

The pH of the exposure media had moderate effects on the uptake rate constants of

318

dissolved Cu (ku,d) (Figures S8d-f and Table 1), while NOM concentration and type more

319

strongly influenced ku,d. At pH 7.0, copper uptake was inhibited as SRHA concentration

320

increased (Figures 3a-c), with ku,d decreasing from (64.6±3.7)×10−6 with no NOM to

321

(11.2±0.6)×10−6 L embryo−1 h−1 with 10 mg-C L−1 SRHA (Table 1). At a constant NOM

322

concentration (5 mg-C L−1), the NOM isolate with the higher SUVA280 value decreased

323

dissolved Cu uptake (Figures 4a-c). Thus, the uptake rate constant of dissolved Cu was

324

dependent upon both NOM concentration and NOM type, with ku,d negatively correlated to

325

[DOC] multiplied by SUVA280 (Figure S12a). This result is consistent with previous findings

326

that NOM with higher aromaticity and SUVA tends to bind more strongly with metal ions (e.g.

327

Zn2+ and Cu2+) and are more effective in decreasing dissolved metal bioavailability.55, 56 This

328

combined parameter ([DOC]×SUVA280) reflects both NOM type and concentration and could be

329

interpreted to represent the concentration of aromatic carbon associated with dissolved NOM.

330



331 332 333 334 335 336 337 338 339 340

Figure 3. Cu uptake in Gulf killifish embryos exposed to dissolved Cu (a, b, and c) or CuO NPs (d, e, and f) with SRHA at different concentrations. Gulf killifish embryos (48 hpf) were exposed to mixtures of 5‰ ASW (~90 mM ionic strength) buffered to pH 7.0 with 2 mM MOPS and containing Cu and NOM combinations as shown. Solid and open symbols represent Cu content measured in whole embryos and chorions, respectively, and dash lines represent modeled uptake kinetics. Circle, square and triangle symbols represent experimental replicates from different batches of embryos (replicate number N = 2 or 3). The Cu contents are averages of 4 embryos/chorions in each sample.


Page 20 of 30

Page 21 of 30


341 342 343 344 345 346 347 348

Figure 4. Cu uptake in Gulf killifish embryos exposed to dissolved Cu (a, b, and c) or CuO NPs (d, e, and f) with different NOM isolates at 5 mg-C L−1. Exposure mixtures comprised of 5‰ ASW (~90 mM ionic strength) buffered to pH 7.0 with 2 mM MOPS, and the following NOM isolate: (a, d) WLHpoA, (b, e) SRFA or (c, f) SRHA. Different symbols correspond to experimental replicates from different batches of embryos (replicate number N = 2 or 3).

349

related to aromatic carbon concentration from the NOM and fairly straightforward to interpret,

350

overall Cu uptake rates involving the CuO NPs yielded more complicated relationships with

351

NOM type and concentration. For example, the overall Cu uptake rates in exposure media spiked

352

with 30 µM CuO NPs were similar among different NOM isolates at 5 mg-C L−1 (Figures 4d-f).

353

These observations were due to NP dissolution that occurred at the same time as uptake and the

354

fact that two forms of Cu (dissolved and nanoparticulate) were bioavailable to the organism.

355

Thus, overall Cu uptake was modeled by discretizing the process to dissolved Cu and

356

nanoparticulate CuO uptake pathways (Equation 5). Values of ku,d from the dissolved Cu-only

357

uptake experiments and the dissolution data (Figure 1, Table 1) were used to estimate NP-

While the overall uptake of copper in the dissolved Cu-only experiments was directly



358

specific uptake rate constant ku,NP. It should be noted that this approach was based on the

359

assumption that the presence of NPs did not affect the uptake kinetics of dissolved metal ions;

360

thus the contributions of NPs and ions could be added in a linear fashion.22 This assumption

361

would be valid if Cu accumulation is low and CuO NPs attached to the surface of embryos do

362

not block entrance to pore canals for ion diffusion or transport. The values of ku,NP were also

363

negatively correlated with [DOC]×SUVA280 (Figure S12b), while ku,NP and ku,d were

364

significantly correlated (Figure S13).

365

The importance of aromatic carbon concentration from NOM for nanoparticle uptake and

366

ku,NP values is not as well-known as it is for dissolved metal uptake rates and ku,d. NOM with

367

greater SUVA values are known to also impart increased colloidal stability for nanoparticle

368

suspensions,48, 54 which was also observed in this study (Figure 2). Thus, ku,NP was decreased for

369

CuO NP suspensions with smaller hydrodynamic diameter, and a possible explanation is that

370

larger CuO NP aggregates can settle on embryo surfaces more easily. Quantitative analysis of the

371

relationship between aggregate size and ku,NP was not feasible, mostly due to the polydispersity

372

of the aggregation data. On the other hand, ζ-potential of CuO NPs in the media shifted

373

negatively as [DOC]×SUVA280 increased (Figures 2c and S7, Table 1). This enhanced

374

electrostatic barrier may also decrease the bioavailability of the CuO NPs, as demonstrated for

375

humic acid-alleviated toxicity of Ag NPs to aquatic organisms of different trophic levels (algae,

376

water fleas and zebrafish larvae).57 In sum, both smaller aggregate size (and thus slower settling

377

velocity) and lower surface affinity (i.e., likely due to increased electrostatic barrier and possibly

378

due to NOM-induced steric hindrance) may have contributed to the decrease in ku,NP at higher

379

aromatic carbon concentration.


Page 22 of 30

Page 23 of 30

380


Relative Contributions of CuO NPs and Dissolved Cu for Cu Uptake. The relative

381

contribution of dissolved Cu uptake and CuO NP uptake in Equation 5 can be described by

382

taking the ratio of the individual terms that contribute to the overall change in Cu concentration

383

in the organism (d[Cu]org/dt) as shown in the following:

384

k u,d [Cu]d Uptake rate dissolved Cu = Uptake rate CuO NP k u,NP ([Cu]T − [Cu]d )

(6)

385

The calculated Cu uptake rate ratio as shown in Equation 6 depends mainly on the extent

386

of CuO NP dissolution (Figure S14). For experiments performed at pH 6.3, and with 5 or 10 mg-

387

C L−1 SRHA, the Cu uptake ratio was >1 most of the time (Figures S14d, h, i). With 5 mg-C L−1

388

SRFA, the ratio increased above 1 after 12 h (Figure S14g), but under the other water chemistry

389

conditions the ratio was < 1. In particular, for experiments with 0.1 mg-C L−1 WLHpoA or at pH

390

7.5, where only 1–2% of the CuO NPs dissolved within the first 12 h, the ratio remained at 0.2–

391

0.3 (Figure S14b, f).

392

This trend was extrapolated over a broad range of [DOC]×SUVA280 values and

393

percentage of CuO NP dissolution (Figure 5), using the ku,d and ku,NP values obtained from the

394

experiments (Figure S12). The relative contributions of dissolved Cu and nanoparticulate CuO to

395

overall Cu uptake rates were mainly dependent upon the percentage of CuO NP that dissolved in

396

the media, and were insensitive to the [DOC]×SUVA280 parameter. The reason for the

397

insensitivity is that the uptake rate constants of both dissolved Cu and nanoparticulate CuO were

398

decreased roughly proportionally with higher aromatic dissolved carbon (Figures S12 and S13).

399

Over the tested [DOC]×SUVA280 range of 0.1–50 m−1, uptake rates of dissolved Cu and

400

nanoparticulate CuO were equal when 4.5−6.1% of the CuO NPs had dissolved. When CuO NPs

401

dissolved to a higher degree (e.g. in response to pH changes or undersaturation conditions58),

402

uptake rate of dissolved Cu will exceed that of nanoparticulate CuO. 23 ACS Paragon Plus Environment


403

404 405 406 407 408 409

Figure 5. Relative contributions of dissolved Cu uptake and CuO NP uptake as influenced by [DOC]×SUVA280 and percentage of CuO NP dissolution. The relative rates of uptake (shown in color) were extrapolated from measurement data (black symbols; N = 81) from all CuO NP exposure experiments.

410

Environmental Implications. The behavior and bioavailability of metal-based NPs and

411

metal ions released from the NPs in aquatic environments depend upon not only the

412

concentration of NOM but also the characteristics of NOM. This study quantitatively

413

demonstrated that NOM with higher SUVA and at higher concentrations decreased the uptake

414

rate constants of both dissolved Cu and nanoparticulate CuO. However, this study also indicated

415

that the ratio of these uptake rate constants did not change with NOM type and concentration.

416

Hence, the bioavailability of CuO NPs is in proportion to that of dissolved Cu under a variety of

417

solution conditions.

418

Aromatic carbon concentration, as indicated by the combined parameter of

419

[DOC]×SUVA280, is a determinant of the bioavailability of both NPs and dissolved metal ions.

420

This parameter [DOC]×SUVA280 is essentially the decadal absorption coefficient at 280 nm (i.e., 24 ACS Paragon Plus Environment

Page 24 of 30

Page 25 of 30


421

a280) of the water matrices, and may be applicable for environmental settings where dissolved

422

NOM is the dominant UV absorbing constituent. Exceptions to this might be waters containing

423

dissolved ferric ions, which can contribute to UV absorbance.59

424

The relative contributions of dissolved Cu and nanoparticulate CuO to overall Cu uptake

425

kinetics are mainly determined by the extent of CuO NP dissolution, which in turn is influenced

426

by multiple environmental and physiological parameters, including pH and organic ligands such

427

as NOM. While it is imperative to continue mechanistic studies investigating the effects of

428

complex water chemistry conditions on the dissolution rates of metal-based nanomaterials, our

429

findings support the use of dissolution rate as one of the critical functional assays for forecasting

430

nanomaterial risk in complex and varied environmental systems.60 Moreover, since the

431

bioavailability of dissolved metal has been extensively investigated, with uptake rates

432

determined under various conditions, findings of this study suggest that existing knowledge on

433

dissolved metal bioavailability and toxicity is instrumental in constructing models for predicting

434

the environmental and human health risks of nanomaterials.

435 436 437

Acknowledgments To our dear friend and mentor George Aiken, who not only helped to unlock the

438

mysteries of dissolved organic matter but also generously taught us the joys of fellowship in

439

scientific discovery.

440

We are also grateful for the insights and assistance from Drs. Marc Deshusses, Amrika

441

Deonarine, Brett Poulin, Joe Ryan, and Keith Lucey, and to Alexis Carey for the image of the

442

Gulf killifish embryo. The work was supported by the National Science Foundation (NSF)

443

(CBET-1066781) and the Center for the Environmental Implications of NanoTechnology (DBI-



444

1266252), which is funded by the NSF and the U.S. Environmental Protection Agency (EPA).

445

Partial support was provided by the C. Gus Glasscock, Jr. Endowed Fund for Excellence in

446

Environmental Sciences at Baylor University. Additional support was provided by the U.S.

447

Geological Survey National Research and Toxic Substances Hydrology Programs. Any use of

448

trade, firm, or product names is for descriptive purposes only and does not imply endorsement by

449

the U.S. Government. This paper has not been subjected to EPA review; therefore, the opinions

450

expressed in this paper are those of the authors and do not necessarily reflect the views of the

451

EPA.

452

Supporting Information Available: Additional information related to the preparation of

453

CuO NP stock suspensions, procedures of the uptake and dissolution experiments, purity and

454

sources of chemicals (Table S1), CuO NP characterization (Figure S1), pH of uptake exposure

455

media (Figure S2), validation of CuO NP dissolution experiments by ASV (Figures S3 and S4),

456

Z-average hydrodynamic diameter (Figure S5) and zeta potential (Figure S7) of CuO NPs, total

457

copper concentration in the media of CuO NP (Figure S6) and dissolved Cu uptake experiments

458

(Figure S11), complete data sets for uptake kinetics of dissolved Cu (Figure S8) and CuO NPs

459

(Figure S9), distribution of Cu in embryos (Figure S10), ku as a function of [DOC]×SUVA280

460

(Figure S12), correlation between ku of dissolved Cu and nanoparticulate CuO (Figure S13), ratio

461

of instantaneous uptake rates of dissolved Cu and CuO NP (Figure S14), and ASV calibration

462

curves (Figure S15). This material is available free of charge via the Internet at

463

http://pubs.acs.org.


Page 26 of 30

Page 27 of 30

464 465 466 467 468 469 470 471 472 473 474 475 476 477 478 479 480 481 482 483 484 485 486 487 488 489 490 491 492 493 494 495 496 497 498 499 500 501 502 503 504 505 506 507 508 509


References 1. Colvin, V. L. The potential environmental impact of engineered nanomaterials. Nat. Biotechnol. 2003, 21, 1166-1170. 2. Wiesner, M. R.; Lowry, G. V.; Alvarez, P.; Dionysiou, D.; Biswas, P. Assessing the risks of manufactured nanomaterials. Environ. Sci. Technol. 2006, 40, 4336-4345. 3. Nel, A.; Xia, T.; Madler, L.; Li, N. Toxic potential of materials at the nanolevel. Science 2006, 311, 622-627. 4. Nowack, B.; Bucheli, T. D. Occurrence, behavior and effects of nanoparticles in the environment. Environ. Pollut. 2007, 150, 5-22. 5. Klaine, S. J.; Alvarez, P. J. J.; Batley, G. E.; Fernandes, T. F.; Handy, R. D.; Lyon, D. Y.; Mahendra, S.; McLaughlin, M. J.; Lead, J. R. Nanomaterials in the environment: Behavior, fate, bioavailability, and effects. Environ. Toxicol. Chem. 2008, 27, 1825-1851. 6. Keller, A. A.; Lazareva, A. Predicted releases of engineered nanomaterials: From global to regional to local. Environ. Sci. Technol. Lett. 2014, 1, 65-70. 7. von Moos, N.; Bowen, P.; Slaveykova, V. I. Bioavailability of inorganic nanoparticles to planktonic bacteria and aquatic microalgae in freshwater. Environ.-Sci. Nano 2014, 1, 214-232. 8. Nel, A. E.; Madler, L.; Velegol, D.; Xia, T.; Hoek, E. M. V.; Somasundaran, P.; Klaessig, F.; Castranova, V.; Thompson, M. Understanding biophysicochemical interactions at the nanobio interface. Nat. Mater. 2009, 8, 543-557. 9. Lowry, G. V.; Gregory, K. B.; Apte, S. C.; Lead, J. R. Transformations of nanomaterials in the environment. Environ. Sci. Technol. 2012, 46, 6893-6899. 10. Misra, S. K.; Dybowska, A.; Berhanu, D.; Luoma, S. N.; Valsami-Jones, E. The complexity of nanoparticle dissolution and its importance in nanotoxicological studies. Sci. Total Environ. 2012, 438, 225-232. 11. Wang, Z. Y.; von dem Bussche, A.; Kabadi, P. K.; Kane, A. B.; Hurt, R. H. Biological and environmental transformations of copper-based nanomaterials. ACS Nano 2013, 7, 87158727. 12. Lin, S. J.; Taylor, A. A.; Ji, Z. X.; Chang, C. H.; Kinsinger, N. M.; Ueng, W.; Walker, S. L.; Nel, A. E. Understanding the transformation, speciation, and hazard potential of copper particles in a model septic tank system using zebrafish to monitor the effluent. ACS Nano 2015, 9, 2038-2048. 13. Huynh, K. A.; McCaffery, J. M.; Chen, K. L. Heteroaggregation reduces antimicrobial activity of silver nanoparticles: Evidence for nanoparticle-cell proximity effects. Environ. Sci. Technol. Lett. 2014, 1, 361-366. 14. Tangaa, S. R.; Selck, H.; Winther-Nielsen, M.; Khan, F. R. Trophic transfer of metalbased nanoparticles in aquatic environments: a review and recommendations for future research focus. Environ.-Sci. Nano 2016, 3, 966-981. 15. Croteau, M. N.; Misra, S. K.; Luoma, S. N.; Valsami-Jones, E. Silver bioaccumulation dynamics in a feshwater invertebrate after aqueous and dietary exposures to nanosized and ionic Ag. Environ. Sci. Technol. 2011, 45, 6600-6607. 16. Bulcke, F.; Thiel, K.; Dringen, R. Uptake and toxicity of copper oxide nanoparticles in cultured primary brain astrocytes. Nanotoxicology 2014, 8, 775-785. 17. Khan, F. R.; Paul, K. B.; Dybowska, A. D.; Valsami-Jones, E.; Lead, J. R.; Stone, V.; Fernandes, T. F. Accumulation dynamics and acute toxicity of silver nanoparticles to Daphnia magna and Lumbriculus variegatus: Implications for metal modeling approaches. Environ. Sci. Technol. 2015, 49, 4389-4397. 27 ACS Paragon Plus Environment


510 511 512 513 514 515 516 517 518 519 520 521 522 523 524 525 526 527 528 529 530 531 532 533 534 535 536 537 538 539 540 541 542 543 544 545 546 547 548 549 550 551 552 553

Page 28 of 30

18. Bohme, S.; Stark, H. J.; Reemtsma, T.; Kuhnel, D. Effect propagation after silver nanoparticle exposure in zebrafish (Danio rerio) embryos: a correlation to internal concentration and distribution patterns. Environ.-Sci. Nano 2015, 2, 603-614. 19. Xiao, Y. L.; Vijver, M. G.; Chen, G. C.; Peijnenburg, W. Toxicity and accumulation of Cu and ZnO nanoparticles in Daphnia magna. Environ. Sci. Technol. 2015, 49, 4657-4664. 20. Heggelund, L. R.; Diez-Ortiz, M.; Lofts, S.; Lahive, E.; Jurkschat, K.; Wojnarowicz, J.; Cedergreen, N.; Spurgeon, D.; Svendsen, C. Soil pH effects on the comparative toxicity of dissolved zinc, non-nano and nano ZnO to the earthworm Eisenia fetida. Nanotoxicology 2014, 8, 559-572. 21. Auffan, M.; Matson, C. W.; Rose, J.; Arnold, M.; Proux, O.; Fayard, B.; Liu, W.; Chaurand, P.; Wiesner, M. R.; Bottero, J. Y.; Di Giulio, R. T. Salinity-dependent silver nanoparticle uptake and transformation by Atlantic killifish (Fundulus heteroclitus) embryos. Nanotoxicology 2014, 8, 167-176. 22. Wang, J.; Wang, W. X. Salinity influences on the uptake of silver nanoparticles and silver nitrate by marine medaka (Oryzias melastigma). Environ. Toxicol. Chem. 2014, 33, 632640. 23. Son, J.; Vavra, J.; Forbes, V. E. Effects of water quality parameters on agglomeration and dissolution of copper oxide nanoparticles (CuO-NPs) using a central composite circumscribed design. Sci. Total Environ. 2015, 521, 183-190. 24. Gondikas, A. P.; Morris, A.; Reinsch, B. C.; Marinakos, S. M.; Lowry, G. V.; Hsu-Kim, H. Cysteine-induced modifications of zero-valent silver nanomaterials: Implications for particle surface chemistry, aggregation, dissolution, and silver speciation. Environ. Sci. Technol. 2012, 46, 7037-7045. 25. Yang, S. P.; Bar-Ilan, O.; Peterson, R. E.; Heideman, W.; Hamers, R. J.; Pedersen, J. A. Influence of humic acid on titanium dioxide nanoparticle toxicity to developing zebrafish. Environ. Sci. Technol. 2013, 47, 4718-4725. 26. 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, 3486-3495. 27. Adeleye, A. S.; Conway, J. R.; Perez, T.; Rutten, P.; Keller, A. A. Influence of extracellular polymeric substances on the long-term fate, dissolution, and speciation of copperbased nanoparticles. Environ. Sci. Technol. 2014, 48, 12561-12568. 28. Chambers, B. A.; Afrooz, A.; Bae, S.; Aich, N.; Katz, L.; Saleh, N. B.; Kirisits, M. J. Effects of chloride and ionic strength on physical morphology, dissolution, and bacterial toxicity of silver nanoparticles. Environ. Sci. Technol. 2014, 48, 761-769. 29. Walsh, M. J.; Goodnow, S. D.; Vezeau, G. E.; Richter, L. V.; Ahner, B. A. Cysteine enhances bioavailability of copper to marine phytoplankton. Environ. Sci. Technol. 2015, 49, 12145-12152. 30. Mudunkotuwa, I. A.; Pettibone, J. M.; Grassian, V. H. Environmental implications of nanoparticle aging in the processing and fate of copper-based nanomaterials. Environ. Sci. Technol. 2012, 46, 7001-7010. 31. Brayner, R.; Ferrari-Iliou, R.; Brivois, N.; Djediat, S.; Benedetti, M. F.; Fievet, F. Toxicological impact studies based on Escherichia coli bacteria in ultrafine ZnO nanoparticles colloidal medium. Nano Lett. 2006, 6, 866-870.


Page 29 of 30

554 555 556 557 558 559 560 561 562 563 564 565 566 567 568 569 570 571 572 573 574 575 576 577 578 579 580 581 582 583 584 585 586 587 588 589 590 591 592 593 594 595 596 597 598


32. Laban, G.; Nies, L. F.; Turco, R. F.; Bickham, J. W.; Sepulveda, M. S. The effects of silver nanoparticles on fathead minnow (Pimephales promelas) embryos. Ecotoxicology 2010, 19, 185-195. 33. Poynton, H. C.; Lazorchak, J. M.; Impellitteri, C. A.; Blalock, B.; Smith, M. E.; Struewing, K.; Unrine, J.; Roose, D. Toxicity and transcriptomic analysis in Hyalella azteca suggests increased exposure and susceptibility of epibenthic organisms to zinc oxide nanoparticles. Environ. Sci. Technol. 2013, 47, 9453-9460. 34. Brun, N. R.; Lenz, M.; Wehrli, B.; Fent, K. Comparative effects of zinc oxide nanoparticles and dissolved zinc on zebrafish embryos and eleuthero-embryos: Importance of zinc ions. Sci. Total Environ. 2014, 476, 657-666. 35. Adam, N.; Leroux, F.; Knapen, D.; Bals, S.; Blust, R. The uptake and elimination of ZnO and CuO nanoparticles in Daphnia magna under chronic exposure scenarios. Water Res. 2015, 68, 249-261. 36. Ribeiro, F.; Gallego-Urrea, J. A.; Goodhead, R. M.; Van Gestel, C. A. M.; Moger, J.; Soares, A.; Loureiro, S. Uptake and elimination kinetics of silver nanoparticles and silver nitrate by Raphidocelis subcapitata: The influence of silver behaviour in solution. Nanotoxicology 2015, 9, 686-695. 37. Luoma, S. N. Bioavailability of trace metals of aquatic organisms - A Review. Sci. Total Environ. 1983, 28, 1-22. 38. Luoma, S. N.; Rainbow, P. S. Why is metal bioaccumulation so variable? Biodynamics as a unifying concept. Environ. Sci. Technol. 2005, 39, 1921-1931. 39. Croteau, M. N.; Misra, S. K.; Luoma, S. N.; Valsami-Jones, E. Bioaccumulation and toxicity of CuO nanoparticles by a freshwater invertebrate after waterborne and dietborne exposures. Environ. Sci. Technol. 2014, 48, 10929-10937. 40. Ramskov, T.; Thit, A.; Croteau, M. N.; Selck, H. Biodynamics of copper oxide nanoparticles and copper ions in an oligochaete - Part I: Relative importance of water and sediment as exposure routes. Aquat. Toxicol. 2015, 164, 81-91. 41. Khan, F. R.; Misra, S. K.; Garcia-Alonso, J.; Smith, B. D.; Strekopytov, S.; Rainbow, P. S.; Luoma, S. N.; Valsami-Jones, E. Bioaccumulation dynamics and modeling in an estuarine invertebrate following aqueous exposure to nanosized and dissolved silver. Environ. Sci. Technol. 2012, 46, 7621-7628. 42. 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, 31963201. 43. Conway, J. R.; Adeleye, A. S.; Gardea-Torresdey, J.; Keller, A. A. Aggregation, dissolution, and transformation of copper nanoparticles in natural waters. Environ. Sci. Technol. 2015, 49, 2749-2756. 44. Kaweeteerawat, C.; Chang, C. H.; Roy, K. R.; Liu, R.; Li, R. B.; Toso, D.; Fischer, H.; Ivask, A.; Ji, Z. X.; Zink, J. I.; Zhou, Z. H.; Chanfreau, G. F.; Telesca, D.; Cohen, Y.; Holden, P. A.; Nel, A. E.; Godwin, H. A. Cu nanoparticles have different impacts in Escherichia coli and Lactobacillus brevis than their microsized and ionic analogues. ACS Nano 2015, 9, 7215-7225. 45. Burnett, K. G.; Bain, L. J.; Baldwin, W. S.; Callard, G. V.; Cohen, S.; Di Giulio, R. T.; Evans, D. H.; Gomez-Chiarri, M.; Hahn, M. E.; Hoover, C. A.; Karchner, S. I.; Katoh, F.; MacLatchy, D. L.; Marshall, W. S.; Meyer, J. N.; Nacci, D. E.; Oleksiak, M. F.; Rees, B. B.; Singer, T. D.; Stegeman, J. J.; Towle, D. W.; Van Veld, P. A.; Vogelbein, W. K.; Whitehead, A.;



599 600 601 602 603 604 605 606 607 608 609 610 611 612 613 614 615 616 617 618 619 620 621 622 623 624 625 626 627 628 629 630 631 632 633 634 635 636 637 638 639 640 641 642 643

Winn, R. N.; Crawford, D. L. Fundulus as the premier teleost model in environmental biology: Opportunities for new insights using genomics. Comp. Biochem. Physiol. D 2007, 2, 257-286. 46. Dubansky, B.; Whitehead, A.; Miller, J. T.; Rice, C. D.; Galvez, F. Multitissue molecular, genomic, and developmental effects of the deepwater horizon oil spill on resident Gulf killifish (Fundulus grandis). Environ. Sci. Technol. 2013, 47, 5074-5082. 47. Jiang, C. J.; Hsu-Kim, H. Direct in situ measurement of dissolved zinc in the presence of zinc oxide nanoparticles using anodic stripping voltammetry. Environ. Sci.-Process Impacts 2014, 16, 2536-2544. 48. Jiang, C. J.; Aiken, G. R.; Hsu-Kim, H. Effects of natural organic matter properties on the dissolution kinetics of zinc oxide nanoparticles. Environ. Sci. Technol. 2015, 49, 11476-11484. 49. Oziolor, E. M.; Bigorgne, E.; Aguilar, L.; Usenko, S.; Matson, C. W. Evolved resistance to PCB- and PAH-induced cardiac teratogenesis, and reduced CYP1A activity in Gulf killifish (Fundulus grandis) populations from the Houston Ship Channel, Texas. Aquat. Toxicol. 2014, 150, 210-219. 50. Gao, Y.; Korshin, G. Effects of NOM properties on copper release from model solid phases. Water Res. 2013, 47, 4843-4852. 51. Miao, A. J.; Zhang, X. Y.; Luo, Z. P.; Chen, C. S.; Chin, W. C.; Santschi, P. H.; Quigg, A. Zinc oxide-engineered nanoparticles: Dissolution and toxicity to marine phytoplankton. Environ. Toxicol. Chem. 2010, 29, 2814-2822. 52. Vencalek, B. E.; Laughton, S. N.; Spielman-Sun, E.; Rodrigues, S. M.; Unrine, J. M.; Lowry, G. V.; Gregory, K. B. In Situ Measurement of CuO and Cu(OH)2 nanoparticle dissolution rates in quiescent freshwater mesocosms. Environ. Sci. Technol. Lett. 2016, 3, 375380. 53. Domingos, R. F.; Tufenkji, N.; Wilkinson, K. J. Aggregation of titanium dioxide nanoparticles: Role of a fulvic acid. Environ. Sci. Technol. 2009, 43, 1282-1286. 54. Deonarine, A.; Lau, B. L. T.; Aiken, G. R.; Ryan, J. N.; Hsu-Kim, H. Effects of humic substances on precipitation and aggregation of zinc sulfide nanoparticles. Environ. Sci. Technol. 2011, 45, 3217-3223. 55. Baken, S.; Degryse, F.; Verheyen, L.; Merckx, R.; Smolders, E. Metal complexation properties of freshwater dissolved organic matter are explained by its aromaticity and by anthropogenic ligands. Environ. Sci. Technol. 2011, 45, 2584-2590. 56. Smith, K. S.; Ranville, J. F.; Lesher, E. K.; Diedrich, D. J.; McKnight, D. M.; Sofield, R. M. Fractionation of fulvic acid by iron and aluminum oxides-Influence on copper toxicity to Ceriodaphnia dubia. Environ. Sci. Technol. 2014, 48, 11934-11943. 57. Wang, Z.; Quik, J. T. K.; Song, L.; Van den Brandhof, E. J.; Wouterse, M.; Peijnenburg, W. Humic substances alleviate the aquatic toxicity of polyvinylpyrrolidone-coated silver nanoparticles to organisms of different trophic levels. Environ. Toxicol. Chem. 2015, 34, 12391245. 58. Kent, R. D.; Vikesland, P. J. Dissolution and persistence of copper-based nanomaterials in undersaturated solutions with respect to cupric solid phases. Environ. Sci. Technol. 2016. 59. Poulin, B. A.; Ryan, J. N.; Aiken, G. R. Effects of iron on optical properties of dissolved organic matter. Environ. Sci. Technol. 2014, 48, 10098-10106. 60. Hendren, C. O.; Lowry, G. V.; Unrine, J. M.; Wiesner, M. R. A functional assay-based strategy for nanomaterial risk forecasting. Sci. Total Environ. 2015, 536, 1029-1037.


Page 30 of 30

Relative Contributions of Copper Oxide ... - ACS Publications

Recommend Documents