Dissolution of Cu-based Engineered Nanomaterials (ENMs) in ...

Dissolution of Cu-based Engineered Nanomaterials (ENMs) in agricultural soil :

Impacts on metal bioavailability and toxicity

Submitted in partial fulfillment of the requirements for the degree of

Doctor of Philosophy in

Department of Civil and Environmental Engineering

Xiaoyu Gao

B.S., Environmental Science, Nanjing University, China

M.S., Civil and Environmental Engineering, Carnegie Mellon University

Carnegie Mellon University Pittsburgh, PA

August 2019

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Acknowledgement Many people that have helped me during my Ph.D. studies in Civil and Environmental Engineering

Department of Carnegie Mellon University. I would like to express my thanks to five groups of

people, without whom this thesis would not have been possible: My advisors, my thesis committee

members, my collaborators, my lab members, funding agencies, and my family.

My Advisors

First, I would like to thank my two advisors, Dr. Gregory V. Lowry, and Dr. Elizabeth Casman. Dr.

Lowry is one of the reasons that I chose CMU as the school to do my graduate level studies. His

research in Environmental Nanotechnology attracted me here. I came to CMU as a Master student,

but I was able to do some interesting independent studies with Dr. Lowry. I had a great time in lab

and research group meeting, and I finally became a Ph.D. student after finishing my MS study.

During the 4 years of Ph.D. study, Dr. Lowry gave me tremendous help in my research. He guided

me to think independently and helped me fix the problem that I faced in my research. Dr. Elizabeth

Casman became my advisor after I passed the qualification exam. She provides tons of insightful

ideas regarding environmental policy and data analysis. Dr. Casman also provided me help on my

scientific writing. Writing ability was a problem when I entered the Ph.D. program because I am not

a native English speaker. With the help from Liz, now I am more confident about scientific writing.

Together with Dr. Lowry and Dr. Liz, we published five very good stories about the environmental

implications of ENMs, and a few more papers will be published soon. I feel truly grateful to be a

Ph.D. student of Dr. Lowry and Dr. Liz. Without their guidance, I could not finish all of the work.

Thesis Committee Members

Besides my advisor, I would like to thank the rest of my dissertation committee members (Dr. Sonia

M. Rodrigues and Dr. David A. Dzombak) for their support and invaluable advice. Dr. Sonia M.

Rodrigues is also our collaborator. She is an expert in soil chemistry and provided lots of valuable

advice with regards to the behavior and bioavailability of metals in the environment. Dr. David A.

Dzombak is our Department head, who taught me Environmental Chemistry during my 1st year in

CMU. His class was still one of my favorite classes, and he really inspired me to do research in this

field. His comments, like the redox reactions of Cu in soil, really helped me to make my studies

more well-rounded.

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My Lab members

I would like to thank my all lab mates, including Yilin Zhang, Garret Bland, Astrid Avellan, Jiang

Xu, Zimo Lou, Eleanor Spielman-Sun, Stephanie Laughton, Eric McGivney, Joe Moore, Rucha

Vaidya, and John Stegemeier. All of them have provided valuable feedback on my studies. Many of

them are very close collaborators for my studies.

I would also like to thank our lab managers, Ronald Ripper and Brian Belowich. My experiments

could not run successfully without their help. I can’t remember how many times that I was very

worried about my experiments and ask them for help. They really helped me so much, so that I

could successfully collect my experimental data.

CEE faculty and staff

I am also grateful to all the CEE staff. They helped me to make sure my Ph.D. studies can go

smoothly. They have also held lots of CEE events to make my Ph.D. life pleasurable.

Other faculty in CEE have also provided me help during my Ph.D. studies. I especially thank Dr.

Mitchell Small for helping me with the modeling and statistical tests in my study. I also thank Dr.

Kelvin Gregory and Dr. Robert Tilton for their comments in my research.

Funding Supports

I thank the NSF and EPA funding under NSF Cooperative Agreement EF-1266252, Center for the

Environmental Implications of NanoTechnology (CEINT), and CBET-1530563 (NanoFARM) for

their funding support.

I thank Carnegie Mellon University, College of Engineering for providing me College of

Engineering Dean’s Fellowship. I also thank Carnegie Mellon University’s Graduate Student

Assembly and Carnegie Mellon University’s Civil and Environmental Engineering Department

(under Professor Steven J. Fenves’s travel grant) for providing me travel funds.

My Family and Friends

Last but not least, I would like to express my deepest thanks to my family and friends. My mother

and my father supported me all the way to my Ph.D. studies. I feel grateful that they provided me a

chance to let me study abroad. I also thank my wife for staying with me and supporting me all the

time. This dissertation would not have been possible without their continued support.

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Table of Contents Acknowledgement .............................................................................................................................................. i

List of Figures ................................................................................................................................................... vii

List of tables ....................................................................................................................................................... x

ABSTRACT ....................................................................................................................................................... xi

CHAPTER 1: Introduction ......................................................................................................................... - 1 -

1.1 Introduction .................................................................................................................................. - 1 -

1.2 Objectives and overview of this thesis ............................................................................................ - 4 -

1.3 References of Chapter 1 .................................................................................................................... - 6 -

CHAPTER 2: Develop a functional assay to measure the dissolution kinetics of metal-based nanoparticles in soil ....................................................................................................................................- 10 -

Abstract: ...................................................................................................................................................- 10 -

2.1 Introduction ......................................................................................................................................- 11 -

2.2 Method and Materials ......................................................................................................................- 13 -

2.2.1 Chemicals ...................................................................................................................................- 13 -

2.2.2Nanoparticle Characterization. ................................................................................................- 13 -

2.2.3 Soils and Characterization of Soil Properties. ......................................................................- 14 -

2.2.4 Soil amendment and incubation. ............................................................................................- 14 -

2.2.5 Total Metal Concentration. .....................................................................................................- 15 -

2.2.6 Extractions to assess the labile Cu in soil samples. .............................................................- 15 -

2.2.7 Determination of Cu speciation in soils. ..............................................................................- 15 -

2.2.8 Dissolution kinetics. .................................................................................................................- 16 -

2.3 Results and Discussion ....................................................................................................................- 16 -

2.3.1 Soil and nanoparticle characterization. ..................................................................................- 16 -

2.3.2 The change of soil pH after amendment. .............................................................................- 17 -

2.3.3 General trends in extractable Cu for CuO NP- and Cu(NO3)2-amended soil. ...............- 18 -

2.3.4 Fractions of dissolved Cu and particulate Cu in extracts. ..................................................- 20 -

2.3.5 Effect of CuO NP concentration on its extractability in soil. ...........................................- 21 -

2.3.6 Dissolution rate of CuO NP in soil. ......................................................................................- 21 -

2.3.7 Effect of aging on speciation of Cu in Cu(NO3)2 and CuO NP amended soil. ..............- 22 -

2.4 Environmental Implications ...........................................................................................................- 24 -

2.5 References of Chapter 2 ..................................................................................................................- 26 -

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CHAPTER 3: Quantify the effect of soil properties, including soil moisture content, organic carbon content and pH, on the dissolution kinetics of CuO NP in soil..........................................................- 30 -

Abstract: ...................................................................................................................................................- 30 -

3.1 Introduction ......................................................................................................................................- 31 -


3.2.1 Chemicals ...................................................................................................................................- 33 -

3.2.2Nanoparticle Characterization .................................................................................................- 34 -

3.2.3 Soil amendment ........................................................................................................................- 34 -

3.2.4 Extraction procedure to measure the fraction of dissolved CuO NP and soil pH. .......- 35 -

3.2.5 Determination of Cu speciation in soils ...............................................................................- 35 -

3.2.6 Dissolution models. .................................................................................................................- 35 -


3.3.1 Effect of Soil Organic Matter on dissolution of CuO NP in soil. ....................................- 38 -

3.3.2 Effect of soil pH on dissolution of CuO NP in soil. ..........................................................- 39 -

3.3.3 Effect of soil moisture content on the dissolution rate and solubility of CuO NP in soil. .. - 40 -

3.3.4 Dissolution rate and solubility of CuO NPs in soils with various properties. ................- 41 -

3.3.5 Predicting CuO NP solubility and dissolution rate in two test soils. ................................- 45 -



CHAPTER 4: Elucidate CuO NP dissolution behavior and toxicity to wheat (Triticum aestivum) in rhizosphere and bulk soil. ..........................................................................................................................- 52 -

Abstract: ...................................................................................................................................................- 52 -

4.1 Introduction ......................................................................................................................................- 53 -


4.2.1 Chemicals ...................................................................................................................................- 55 -

4.2.2Nanoparticle Characterization .................................................................................................- 56 -

4.2.3 Soils and Characterization of Soil Properties .......................................................................- 56 -

4.2.4 Soil amendment. .......................................................................................................................- 56 -

4.2.5 Germination and plant growth ...............................................................................................- 57 -

4.2.6 Sampling of soil and plant tissue ............................................................................................- 57 -

4.2.7 Soil extraction ...........................................................................................................................- 58 -

4.2.8 Cytoviva analysis .......................................................................................................................- 58 -


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4.3.1 Nanoparticle characterization .................................................................................................- 59 -

4.3.2 Change in extractability of Cu in bulk soil during the plant growth experiment. ............- 59 -

4.3.3 Toxicity of CuSO4 and CuO NP............................................................................................- 61 -

4.3.4 Cu root association and Cu uptake ........................................................................................- 62 -

4.3.5 Effect of near-root environment on Cu availability from CuO NP treatment ...............- 64 -

4.3.6 Soil pH in bulk soil, rhizosphere soil and loosely attached soil .........................................- 65 -

4.4 Discussion .........................................................................................................................................- 67 -

4.4.1 CuO NP dissolution is linked to toxicity. .............................................................................- 67 -

4.4.2 CaCl2 extractable Cu correlates with toxicity of CuO NP to wheat. ................................- 68 -

4.4.3 Root-associated CuO NP modulates toxicity. .....................................................................- 68 -

4.4.4 Root exudates affect CuO NP dissolution and availability. ...............................................- 68 -

4.4.5 Triticum aestivum regulated Cu uptake. ....................................................................................- 70 -

4.5 Agricultural implications .................................................................................................................- 70 -


CHAPTER 5: Dissolution functional assay improves understanding of metallic nanoparticle toxicity in agricultural soil ........................................................................................................................................- 76 -

Abstract ....................................................................................................................................................- 76 -

5.1 Introduction ......................................................................................................................................- 77 -

5.2 Methods .............................................................................................................................................- 77 -

5.2.1 Dissolution profile measurement assay .................................................................................- 77 -

5.2.2 Plant uptake measurement and toxicity ................................................................................- 78 -

5.3 Results and discussion: ....................................................................................................................- 78 -

5.3.1 Differences in dissolution time scale require different assays ...........................................- 78 -

5.3.2 Dissolution profile measurement assay predicted toxicity of Cu species to Triticum aestivum ..............................................................................................................................................................- 79 -


5.4 References for Chapter 5 ................................................................................................................- 81 -

CHAPTER 6: Summary of Major Contributions and Perspective on Future Research ..................- 83 -

6.1 Summary of Major Contribution ...................................................................................................- 83 -

6.1.1. Major Contribution from Objective 1: A test method to measure dissolution of CuO NP in soil was developed. ........................................................................................................................- 83 -

6.1.2. Major Contribution from Objective 2: A model was developed to evaluate the effect of soil pH and organic carbon content on dissolution kinetics of CuO NP in soil. ....................- 83 -

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6.1.3. Major Contributions from Objective 3:Dissolution of CuO NPs in soil was correlated with its toxicity to wheat (Triticum aestivum). Dissolution of CuO NPs under the influence of root activity in rhizosphere soil was quantified. ............................................................................- 84 -

6.1.4. Major Contribution from Objective 4: Dissolution kinetics functional assays were used to estimate exposure to ionic Cu from Cu-based ENMs in soil. This exposure correlated to observed toxicity in wheat. ...............................................................................................................- 84 -

6.2 Perspectives for future research .....................................................................................................- 85 -

6.2.1. Extension of the model to predict the behavior of other metal/metal oxide ENMs in soil ..............................................................................................................................................................- 85 -

6.2.2. Optimize the way to measure toxicity of metal/metal oxide ENMs in soil. ..................- 88 -

6.2.3. Design ENMs that can solve the micronutrient deficiency problem in calcareous soil. - 88 -

6.3 References for Chapter 6 ................................................................................................................- 89 -

Appendices ..............................................................................................................................................- 92 -

Appendix 1- Supporting information for Chapter 2: Develop a functional assay to measure the dissolution kinetics of metal-based nanoparticles in soil. .................................................................- 92 -

Appendix 2- Supporting information for chapter 3:Quantify the effect of soil properties, including soil moisture content, organic carbon content and pH, on the dissolution kinetics of CuO NP in soil. ......................................................................................................................................................... - 103 -

Appendix 3- Supporting information for Chapter 4: Elucidate CuO NP dissolution behavior and toxicity to wheat (Triticum aestivum) in rhizosphere and bulk soil. ................................................ - 115 -

Appendix 4. Explanation on the solubility of CuO NP in Chapter 3 ......................................... - 125 -

Reference for Appendices: ................................................................................................................. - 126 -

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List of Figures Figure 2-1. Extractable Cu and in CuO NP and Cu(NO3)2 amended soils as a function of time and

the first order dissolution fit for CuO NP in soil. ---------------------------------------------------(19)

Figure 2-2. Fraction of small particles and dissolved ions (those passing 3kDa filter) in (a) DTPA

extracts and (b) CaCl2 extracts. -----------------------------------------------------------------------------(20)

Figure 2-3. Change of Cu speciation in amended soils as inferred by XANES.-------------------(23)

Figure 3-1. Schematic of CuO NP dissolution model. -------------------------------------------------(36)

Figure 3-2. Dissolution kinetics of CuO NP in Lufa 2.1 soil without added SOM (100 mg/kg dw

CuO NP treatment, circles) or with added SOM (300mg/kg dw CuO NP treatment, triangles). (39)

Figure 3-3. DTPA extractable Cu in Lufa 2.2 soil dosed with 500 mg/kg CuO NP at pH 5.9 and

pH 6.8. ------------------------------------------------------------------------------------------------------------(40)

Figure 3-4. Effect of moisture content on the dissolution kinetics of CuO NP in soil. -----------(41)

Figure 3-5. Correlation between organic carbon content and solubility (a) and between {H+}and

dissolution rate constant, kd (b). -------------------------------------------------------------------------------(44)

Figure 3-6. Prediction and experimental data of CuO NP dissolution in an Arizona soil (a) and in a

Portugal soil (b). --------------------------------------------------------------------------------------------------(46)

Figure 4-1. Change in DTPA extractable Cu over time for each treatment: a) CuO NP treatment, b)

CuSO4 treatment, and comparison of mean of extractable Cu for each Cu treatments at the end of

the plant growth period: c) DTPA extraction, d) CaCl2 extraction. ------------------------------------(60)

Figure 4-2. a) Root compactness and b) leaf length (leaf growth stage is noted with number, from 1

being the youngest to 3 the oldest) of wheat seedlings grown in freshly amended and aged CuO NP,

CuSO4-amended soil, and control treatments. -------------------------------------------------------------(62)

Figure 4-3. Hyperspectral imaging of plant roots grown in soil with freshly amended CuO NPs (a-e)

or after aging (f-i). ------------------------------------------------------------------------------------------------(63)

Figure 4-4. CaCl2 and DTPA extractable Cu in fresh (left side) and aged (right side) CuO NP

amended rhizosphere soil, loosely attached soil and bulk soils. -----------------------------------------(65)

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Figure 4-5. Mean ± SD of soil pH (measured using CaCl2 extraction) in rhizosphere soil, loosely

attached soil and bulk soil in a) soil freshly amended with CuO NP, b) aged CuO NP treatment c)

control soil, and; d) Comparison of pH of bulk soil among all treatments.--------------------------(66)

Figure 5-1. The dissolution profile of 250mg/kg of CuO NP, Cu(OH)2 NP and CuSO4 in Lufa 2.2

soil. ----------------------------------------------------------------------------------------------------------------(79)

Figure 5-2. Correlations between (a) Cu2+ integrated exposure and toxicity to Triticum aestivum and

(b) Cu2+ concentration at the end of exposure period and toxicity to Triticum aestivum ------------(80)

Figure A1-1. A) Primary particle size distribution determined from counting primary particles from

10 TEM imagines. B-K) Ten TEM images of CuO NP.--------------------------------------------------(95)

Figure A1-2. Size distribution of 80mg/kg CuO NP in pH=7, 5mM NaHCO3 buffer determined by

dynamic light scattering: (a) Number averaged size distribution, (b) intensity averaged size

distribution and (c) and autocorrelation function-----------------------------------------------------------(96)

Figure A1-3. Zeta potential of 80mg/kg CuO NP as a function of pH measured in (a) 5mM

NaHCO3 buffer and (b) 5mM NaNO3. Error bars indicate ± 1standard error. -----------------------(97)

Figure A1-4. X-ray diffraction spectrum of CuO NP. The CuO NPs used here are identified as

tenorite. ------------------------------------------------------------------------------------------------------------(98)

Figure A1-5. pH of CaCl2 extracts in different amended and blank soils-----------------------------(99)

Figure A1-6. Extractable Cu and in wet and air dried amended soils as a function of time: (a)DTPA

extraction for 10 mg/kg amendment, (b) CaCl2 extraction for 10 mg/kg amendment,(c) DTPA

extraction for 100 mg/kg amendment and (d) CaCl2 extraction for 100 mg/kg amendment. (101)

Figure A1-7. XANES spectra for model compounds----------------------------------------------------(102)

Figure A2-1. Cu EXAFS spectra (black) and linear combination fits (red) for CuO NP and CuSO4

exposed soil. (a) Arizona soil, (b) Lufa 2.2 soil. -----------------------------------------------------------(111)

Figure A2-2. DTPA extractable Cu in Lufa 2.1 soils dosed with 100mg/kg CuO NPs at pH 5.0

(squares) and pH 7.4 (triangles). Bars are standard deviation of the measurements. ----------------(112)

Figure A2-3: Cross validation of the correlation between kd and {H+}. -----------------------------(113)

Figure A3-1. Change in DTPA extractable Cu and CaCl2 extractable Cu for 250mg/kg CuO NP

treatment, 250mg/kg and 500mg/kg CuSO4 treatments (without growing plants) over 30 days. (116)

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Figure A3-2. Different soil regions defined in Chapter 4. ----------------------------------------------(117)

Figure A3-3. (A) Spectral library of the CuO-NPs. The spectral library has been built using

datacubes of CuO mixed with hydrated soil. (B) Example of SAM (Spectral Angle Mapping) results

to test for the specificity of the spectral library using positive controls (soil containing CuO NPs) or

negative controls (soil without CuO NPs or control root) images. The pixels containing the spectral

signal of CuO NP are highlighted in red (bottom line).---------------------------------------------------(120)

Figure A3-4. DTPA extractable Cu on bulk soil and bulk bottom soil in different Cu treatments

In all treatments, no significant differences (P<0.05, unpaired t-test) were found between DTPA

extractable Cu in bulk soil and bulk bottom soil, suggesting no vertical transport of Cu in all

treatments. ------------------------------------------------------------------------------------------------------(121)

Figure A3-5. Representative photos showing Cu toxicity led to shortened root and/or root

compactness in fresh CuSO4 treatment---------------------------------------------------------------------(123)

Figure A3-6. Hyperspectral imaging of plant roots grown in soil with CuO-NP, CuSO4 or Na2SO4

(control) freshly amended or after aging. Roots exposed to CuSO4 (both after soil aging or not)

showed a brown-damaged (necrotic) zone, that was not found on any of the CuO NP exposed

roots. -------------------------------------------------------------------------------------------------------------(124)

Figure A3-7. Mean concentration of Cu (mg/kg) in wheat tissue (dry weight): a) Cu concentration

in shoots, b) Cu concentration in roots. -------------------------------------------------------------------(125)

Figure A4-1. Conceptual of Cu speciation in Lufa 2.2 soil a) CuO NP treatment and b) CuSO4

treatment. -------------------------------------------------------------------------------------------------------(126)

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List of tables Table 2-1. Modeled first-order dissolution parameters for CuO NP amended soil. ---------------(22)

Table 3-1. Dissolution rate and solubility of CuO NP in a range of soils with various properties(42)

Table A1-1. Calibration ranges used for ICP-MS measurement used in Chapter 2 ------------------(93)

Table A1-2. Total Cu measured in amended soils (4 replicates) -----------------------------------------(94)

Table A1-3. Extractable Cu in unamended soil (3 replicates) ----------------------------------------------(94)

Table A1-4. The results of the LCF analysis of the X-Ray Absorption Near Edge Structure

(XANES) region of the samples------------------------------------------------------------------------------(103)

Table A2-1: Properties of sampled soils--------------------------------------------------------------------(106)

Table A2-2: Mass balance and experimental conditions for each treatment-------------------------(106)

Table A2-3. Comparison between CuO NP dissolution measured by XANES and chemical

extraction. --------------------------------------------------------------------------------------------------------(109)

Table A2-4. Linear combination fitting results of k3-weighted Cu EXAFS spectra (Figure S3-1) for

Arizona soil exposed to 300mg/kg of CuO NP or CuSO4. --------------------------------------------(110)

Table A2-5.Multivariate regression between dissolution rate constant and soil organic matter content and hydrogen ion activity. -------------------------------------------------------------------------(114)

Table A2-6.Multivariate regression between solubility and soil organic matter content and hydrogen ion activity ------------------------------------------------------------------------------------------------------(114)

Table A2-7.Multivariate regression between reprecipitation rate constant and soil organic matter content and hydrogen ion activity. --------------------------------------------------------------------------(114)

Table A3-1: Total Cu concentration (mean (SD), mg/kg) in soil for each treatment--------------(115)

Table A3-2: DTPA extractable Cu (mean (SD), mg/kg) in the control treatment before and after plant growth ----------------------------------------------------------------------------------------------------(115)

Table A3-3: Samples that provided sufficient soil for DTPA extraction for rhizosphere soil and loosely attached soil--------------------------------------------------------------------------------------------(122)

Table A3-4: Samples that provided sufficient soils for CaCl2 extraction for rhizosphere soil and loosely attached soil--------------------------------------------------------------------------------------------(122)

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ABSTRACT Metal and metal oxide (Me/MeO) engineered nanomaterials (ENMs) are being used in

agriculture as fertilizers and fungicides. A better understanding of how ENMs behave in agricultural

soil, interact with plants (through both soil application and foliar application), and become

bioavailable to plants can guide us to design safer, and more sustainable ENM enabled

agrochemicals. To that end, this thesis aims at better understanding the fate and behavior of ENMs

in agricultural soil and their bioavailability to plants.

The first objective of this work was to develop a method to measure the dissolution kinetics

of CuO NP in soil. Chemical extractions, CaCl2 and Diethylenetriamine Pentaacetic Acid extraction

over time have been developed to achieve this goal. This method was then applied in the second

objective to investigate how soil properties (pH, moisture content and organic carbon content)

influence the dissolution of CuO NPs in different standard soils and a natural soil. In this study, the

solubility of CuO NP was found to correlate well with soil organic matter content (R2 = 0.89),

independent of soil pH. In contrast, the dissolution rate constant of CuO NP in soil correlated with

soil pH for pH<6.3 (R2 = 0.89), independent of soil organic matter. Moisture content, on the other

hand, showed no impact on the dissolution kinetics of CuO NP in soil. These relationships

predicted the solubility and dissolution rate constants of CuO NP in two non-standard test soils

(pH=5 and pH=7.6).

The third objective was to investigate the bioavailability of CuO NP to plants, and their

effects compared to Cu salts. The third study quantified the influence of time and near-root

chemical conditions on dissolution of CuO NP to investigate influence of such dissolution on its

toxicity to Triticum aestivum. Readily available Cu (as reflected by CaCl2 extraction) increased in

rhizosphere soil, whereas the overall dissolution of CuO NP (as reflected by DTPA extraction)

decreased in rhizosphere soil. On the other hand, aging of CuO NPs increased the toxicity to

Triticum aestivum (reduction in root maximal length).This study stressed the importance of CuO NP

dissolution on its toxicity, and showed that plant-induced changes in rhizosphere conditions should

be considered when measuring the dissolution of CuO NP near roots.

The fourth objective was to develop a functional assay that used CuP NP dissolution

kinetics to estimate the Cu2+ ion exposure, and to determine if this assay can predict toxicity to

plants and to soil organisms. For different Cu based ENMs, dissolved Cu was plotted over time to

get the dissolution profile. Different Cu-based ENM had distinct dissolution rates. The integrated

xii

Cu2+ exposure (the area under the dissolution curve in the dissolution profile) was correlated with

selected biological endpoints. The integrated exposure of CuO NP correlated well with its toxicity to

Triticum aestivum. This study suggested that the dissolution profile of Cu-based ENMs may be a

better measure of exposure to Cu ion than a single measurement of extractable Cu2+ ion at the end

of an experiment.

Overall, meeting these objectives provided fundamental knowledge on how different ENMs

behave in soils, and how they interact with plants. It also provides a test method to measure the

dissolution rate of Cu-based ENMs directly in soils. These methods and knowledge will provide

guidance on the design and application of nano-enabled agrochemicals for improving the

sustainability of agriculture, especially in marginal soils where agriculture productivity is typically

low.

- 1 -

CHAPTER 1: Introduction 1.1 Introduction Metal and metal oxide (Me/MeO) engineered nanomaterials (ENMs) are being used in fertilizers

(e.g. ZnDDP ® or CuDDP®) and fungicides (e.g. Kocide ® 3000) in agriculture1–3. The small size of

nanoparticles (NPs) facilitates their uptake by plants through leaves and roots4,5. Also, they can be

tuned to slowly release active ingredients over time or to target the relevant plant tissue6,7 by

modifying their surface coatings. The overuse of non-nano metal-containing agrochemicals has

resulted in toxicity to plants8, microbial communities9 and rhizosphere and soil invertebrates10. If

properly tuned, nanoparticles could deliver metals more efficiently to plants than non-nano

agrichemicals, reducing the attendant environmental contamination. But, to fulfil the promise of

nano-enabled agrichemicals, we need to understand their behavior in soil, particularly the factors

that influence their dissolution, because information on ENM dissolution in soil will be necessary

for developing guidelines on the appropriate dosage of nano-enabled agrichemicals and for

predicting their impact on human health and the environment.

Previous studies have investigated the toxicity of metal and metal and metal oxide (Me/MeO)

ENMs in aqueous systems. These studies attributed the toxicity of Ag NPs11–13, ZnO NPs14 and

CuO NPs15,16 to the ions released upon their dissolution. By separating the dissolved ions from the

NPs using either filtration or centrifugation methods, previous studies have shown that the

dissolution of Me/MeO ENMs in aqueous systems generally follows a first-order process17–19.

It is more difficult to measure the dissolution of Me/MeO ENMs in soil. In soils, metal ions

released from Me/MeO ENMs associate with soil solid surfaces, e.g. soil organic matter (SOM) and

clay20,21. In soil, the released metal ions need to be chemically extracted from soil in order to measure

the dissolution kinetics.

The lack of characterization of dissolution behavior of Me/MeO ENMs in soil has resulted in

contradictory conclusions on whether the toxicity of Me/MeO ENMs in soil is due to metal ions or

due to a particle-specific effect. Some studies attributed the toxic effect of Me/MeO ENMs in soil

to ion release (dissolution) 22–24, while others concluded the opposite25,26. A few examples that

illustrates the problem follow.

- 2 -

Servin et al. assumed that 10% of the CuO NP would dissolve in soil, the same fraction that

dissolved in pure sand, and used that estimate as the Cu2+ concentration of their ionic control, rather

than measuring actual CuO dissolution in soil. They concluded that dissolution of CuO NPs could

not fully explain the plant toxicity because the plant responses differed from those of their Cu2+ ion

control .25 Much more than 10% CuO NP could have dissolved in soil, compared to sand, because

SOM is a Cu sink that increases the amount of CuO NP that can be dissolved27. Not measuring

dissolution in soil undermined their conclusions about a NP-specific effect. Similar problems

occurred in other studies 22,23,28–30.

Dimkpa et al. (2013) evaluated the total CuO NP dissolved in soil using a water-extraction method. 31 The water-extraction method does not extract Cu bound to the soil solid matrix (which includes

SOM), the location of >90% of the dissolved Cu in soils32–34. Because they underestimated the

actual dissolved Cu in the soil, their conclusion of CuO NP-specific toxicity in soil was false. (There

are other problems with the water extraction method that are discussed later, in Chapter 4.)

Qiu et al. observed that the toxicity of CuO NP, CuO bulk particles and soluble Cu (Cu(AC)2)

depends on their solubility in soil, and that the distinction between NP and bulk particles diminished

after a 90-day aging period. However, the dissolution profile over time (a graph of changing metal

ion concentration over time during an exposure experiment) was not provided (they measured

dissolution only on day 0 and day 90). They successfully correlated the toxicity of CuO NP to roots

of Hordeum vulgare L. (in a 5-day root elongation experiment) with free Cu ions in soil pore water

measured at a single time point before seeding.24 This is a major step in the right direction, but an

even better correlation could have been obtained if dissolution were measured over the 5-day

exposure period. Had the dissolution profile been known, the ionic control could have reflected the

change in Cu2+ over time. This would be necessary to determine whether a NP specific effect

existed. If the dissolution of metal or metal oxide ENMs in soil is not tracked, the attribution of an

observed toxic effect to the NPs or the ion-release process is usually inconclusive.

Chemical extraction methods have been developed to measure the speciation and bioavailability of

metals in soil. The chemical extraction methods can be generally classified as two types: (1) the pore

water extraction method, and (2) the labile metal extraction method. The first method extracts free

mobile metal ions in soil pore water. Such species are usually considered to be readily available to

plants. This method uses a dilute salt as extractant, such as 0.01M calcium chloride (CaCl2) or

0.005M calcium nitrate (Ca(NO3)2). Such extractions predict the bioavailability of metals by

- 3 -

mimicking the chemistry of soil pore water (e.g. similar ionic strength) and targeting the

exchangeable metal ions in soil pore water 35–37. The second method uses strong chelators, such as

0.05M diethylenetriamine pentaacetic acid (DTPA) or 0.05M ethylenediaminetetraacetic acid

(EDTA) as an extracting agent. Those strong chelating agents mimic the chelating effect of root

exudates which enhance the availability of nutrients from the soil rhizosphere for plant uptake38.

This extraction method targets the total “labile” metal in soil that not only includes the free metal

ions in soil pore water, but also the metal ions associated with SOM. In this thesis I show how these

two extraction methods, with modifications, can be used to measure the released ions from ENMs

directly in soil. The behavior of metal ions in soil is a rapid partitioning process between soil pore

water and soil solid surfaces.21 In contrast, NPs slowly releases metal ions into the soil porewater39.

Essentially, this makes the bioavailability of Me/MeO ENMs time-dependent. Thus, to capture the

dissolution kinetics of Me/MeO NPs in soil, chemical extractions need to be performed repeatedly

over time.

When measuring the dissolution of ENMs in soil, the time scale of the dissolution process could

also be system-dependent. In aqueous systems, studies have shown that the main factors affecting

the dissolution kinetics of Me/MeO ENMs include pH,17,40,41 organic matter (OM) content1,42,43,

oxygen concentration44 (if the dissolution is a oxidative dissolution, e.g. AgNPs) and complexing

ions (e.g. Cl-)45. The dissolution of Me/MeO NPs in soil systems is expected to be affected by the

soil properties, e.g. soil pore water pH and organic matter content. These parameters affect either

the ion release rate or the partitioning of the metal ions with soil solid surfaces. It is the partitioning

in this multiphase system (soil, water, ENM) that makes the time scale of dissolution in soil different

from that in water.

The time scale of dissolution and the time scale of exposure are both important when trying to

correlate the dissolution of ENMs in environmental media to biological endpoints

(toxicity/bioavailability). In aqueous systems, the time scale of dissolution is usually much shorter

than the time scale of exposure. A measurement of (equilibrium) solubility of the Me/MeO ENMs

in the relevant aqueous medium should correlate with NP toxicity24,46. However, in soil systems,

where the dissolution kinetics are more complex 47,48, a measurement of the overall dissolution

profile is likely needed to establish the connection between dissolution of ENMs and their biological

endpoints.

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In several regions around the world, soils are deficient in copper, especially regions with calcareous

soils49. There, additional Cu is added to soil to supply the micronutrient for optimal plant growth.

Copper-based ENMs can be applied to soil as a ‘slow-release’ source to deliver Cu to plants. The

Cu-based ENMs have a higher efficiency and lower risk compared to Cu ion based fertilizers.27,50,51.

The other reason that Cu-based ENMs are applied to agricultural soil is due to their anti-microbial

properties52, e.g. Kocide ® 3000 (whose main component is Cu(OH)2 NPs) is a registered fungicide

in U.S. agriculture. Although the agricultural application of Cu-based fertilizers and pesticides are

regulated (maximal application rate is 75 kg ha-1 yr-1) , the regulations were based on toxicity of

soluble species such as CuSO4 or Cu(NO3)227. However, the fate and behavior of Cu based ENMs is

different from the Cu soluble species27,39 leading to potentially different bioavailability and toxicity to

plants and soil organisms compared to more soluble species. Questions remain about how to

accurately measure the toxicity of Cu-based ENMs in soil and about how they should be regulated.

Thus, Cu-based ENMs, especially CuO NP was selected as the main Me/MeO ENMs to investigate

in this thesis. Specifically, in this thesis, dissolution of Cu-based ENMs, which release the active

ingredient- Cu2+, is considered the main transformation process that affect its bioavailability in soil.

Other transformation processes, e.g. redox reactions, were not considered. For aerated agricultural

soil (topsoil to which agrochemicals are applied), the redox potential is usually above 400mv53. At

such a redox condition, Cu(II) is the major Cu valence state54.

1.2 Objectives and overview of this thesis

In order to tackle the challenge of understanding and quantifying metallic ENMs' behavior in

aerobic soil and plant systems, four objectives were pursued, each comprising a separate chapter of

this thesis. Objective 1 was to develop a method for quantifying the dissolution kinetics of CuO

NPs in soil. Objectives 2 and 3 used the new dissolution measurement assay to investigate 1) how

soil properties affect the dissolution behavior of CuO NPs in soil (Objective 2) and how plant

exudates affect the dissolution of CuO NP in soil, and how dissolution of CuO NP affects the plant

health (Objective 3). Objective 4 was to determine the best assay of Cu-based ENM dissolution in

soil to predict their toxicity to plants and soil isopods. This knowledge will provide guidance on the

design and application of nano-enabled agrochemicals for improving the sustainability of agriculture,

especially in marginal soils where agriculture productivity is typically low.

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Objective 1: Develop a functional assay to measure the dissolution kinetics of metal-based

nanoparticles in soil.

I developed a protocol for a time-dependent series of extractions to determine the CuO NP

dissolution rate constant and reveal the dissolution profile in soils. This work has been published in

Environmental Science & Technology. 27

Objective 2: Quantify the effect of soil properties, including soil moisture content, organic

carbon content and pH, on the dissolution kinetics of CuO NP in soil.

Chemical extractions were applied to measure CuO NP dissolution kinetics in soils with different

properties, providing data for a model to predict the dissolution kinetics of CuO NPs in soil. The

model successfully predicted the dissolution kinetics of CuO NPs in two unknown soils. This study

showed that soil pH and organic matter content affect the dissolution behavior of CuO NP in soil in

a predictable manner. This work has been published in Environmental Science & Technology55.

Objective 3: Elucidate CuO NP dissolution behavior and toxicity to wheat (Triticum

aestivum) in rhizosphere and bulk soil.

To quantify the influence of time and near-root chemical conditions on dissolution and lability of

CuO NPs in rhizosphere soil, and to determine the influence of this dissolution on the toxicity of

CuO NPs to plants, we measured the rate of dissolution of CuO NPs in bulk soil, and in soil in

which wheat plants (Triticum aestivum) were grown. At the end of the plant growth period (14 days),

available Cu was measured in three different soil compartments: bulk (not associated with roots),

loosely attached to roots, and rhizosphere (soil firmly attached to roots). Root length shoot length

and biomass were also measured as indicators of toxicity. This study correlated CuO NP dissolution

and the resulting Cu ion exposure profile to phytotoxicity and showed that plant-induced changes in

rhizosphere conditions are the most important determinants of ENM toxicity to roots. This work

has been published in Environmental Science & Technology47.

Objective 4: Use the dissolution kinetics functional assays to predict Cu-based ENMs

toxicity in agricultural soil.

Dissolution profiles of various Cu-based ENMs in Lufa 2.2 soil were measured. Toxicity of different

Cu species to Triticum aestivum (as evidenced by shortened leaves length and root length, reduction in

biomass) was measured. The preliminary analysis of this approach suggested that dissolution profiles

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of different Cu species could be predictors of the different biological endpoints than the status quo

of measuring extractable Cu only at the end of the exposure period.

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(24) Qiu, H.; Smolders, E. Nanospecific phytotoxicity of CuO nanoparticles in soils disappeared when bioavailability factors were considered. Environ. Sci. Technol. 2017.

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(28) Anderson, A.; McLean, J.; McManus, P.; Britt, D. Soil chemistry influences the phytotoxicity of metal oxide nanoparticles. Int. J. Nanotechnol. 2017, 14 (1–6), 15–21.

(29) Watson, J.-L.; Fang, T.; Dimkpa, C. O.; Britt, D. W.; McLean, J. E.; Jacobson, A.; Anderson, A. J. The phytotoxicity of ZnO nanoparticles on wheat varies with soil properties. Biometals 2015, 28 (1), 101–112.

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(39) McShane, H. V. A.; Sunahara, G. I.; Whalen, J. K.; Hendershot, W. H. Differences in soil solution chemistry between soils amended with nanosized CuO or Cu reference materials: implications for nanotoxicity tests. Env. Sci Technol 2014, 48 (14), 8135–8142.

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(43) Jiang, C.; Castellon, B. T.; Matson, C. W.; Aiken, G. R.; Hsu-Kim, H. Relative Contributions of Copper Oxide Nanoparticles and Dissolved Copper to Cu Uptake Kinetics of Gulf Killifish (Fundulus grandis) Embryos. Environ. Sci. Technol. 2017, 51 (3), 1395–1404.

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(48) Gao, X.; Spielman-Sun, E.; Rodrigues, S. M.; Casman, E. A.; Lowry, G. V. Time and Nanoparticle Concentration Affect the Extractability of Cu from CuO NP-Amended Soil. Environ. Sci. Technol. 2017, 51 (4).

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CHAPTER 2: Develop a functional assay to measure the dissolution kinetics of metal-based nanoparticles in soil

Abstract: The effect of CuO nanoparticle (NP) concentration and soil aging time on the extractability of Cu

from a standard aerobic sandy soil (Lufa 2.1) was investigated. The soil was dosed with CuO NP or

Cu(NO3)2 at 10 mg Cu kg-1 soil (mg/kg) or 100 mg/kg total copper, then extracted using either

0.01M CaCl2 or 0.005M DTPA (pH 7.6) extraction fluids at selected times over 31 days. For 100

mg/kg CuO NP, the amount of DTPA-extractable Cu in soil increased from 3 wt% immediately

after mixing to 38 wt% after 31 days. In contrast, the extractability of Cu(NO3)2 was highest initially,

decreasing with time. The increase in extractability was attributed to CuO NP dissolution in soil.

This was confirmed with synchrotron X-ray absorption near edge structure (XANES)

measurements. The CuO NP dissolution kinetics were modeled by a first-order dissolution model.

Our findings indicate that dissolution, concentration, and aging time are important factors

influencing Cu extractability in CuO NP-amended soil, and suggest that a time dependent series of

extractions could be developed as a functional assay to determine the dissolution rate constant.

This work has been published in Environmental Science & Technology as ‘Time and Nanoparticle Concentration Affect the Extractability of Cu from CuO NP-Amended Soil’ , doi: acs.est.6b04705

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2.1 Introduction Copper based nanoparticles (NP) including metallic copper (Cu NP), copper oxides (Cu2O NP and

CuO NP), and copper hydroxides (Cu(OH)2 NP) are manufactured nanomaterials that have been used

as pesticides and fungicides because of their antimicrobial properties1, 2 .They can also be used as

fertilizers to deliver micronutrient-Cu to plants, which can improve fertilizer efficiency and crop

yield3,4. Copper salt (mainly as Cu(NO3)2 or CuSO4) based micronutrients and pesticides have

historically been widely used. Excessive use of Cu containing fertilizers and pesticides may lead to

negative impacts on ecosystems, soil microorganisms, microbial processes5, plants6 and soil

invertebrates7.

In the U.S., Cu containing fertilizers and pesticides are regulated, with the maximum

application rate of 75 kg/ha/year (USEPA, 1993). However, these regulations were determined

using highly soluble Cu salts (e.g. Cu(NO3)2 and CuSO4) in soil. Dynamic processes including

aggregation, oxidation, and dissolution will likely make the available pool of Cu derived from Cu

based NP time-dependent8, 9. While the importance of time on the fate and bioavailability of Cu salts

is documented10-12, aging effects for Cu based NP has not been elucidated. In order to assess the

impact of Cu based NP to agroecosystems, it is important to determine the factors controlling their

bioavailability in soils.

Chemical extraction methods are used to predict the bioavailability of metal in soil13. Several

single extraction methods, originally developed to determine the fraction of metals in soil involved

in geochemical equilibrium processes including sorption and precipitation, can predict the leaching

of soil metals to groundwater, their impact on ecosystems, and their bioavailability for soil organisms

or plants11-22. Two extraction methods, 0.01M CaCl2 extraction and 0.005M

diethylenetriaminepentaacetic acid (DTPA) extraction (pH 7.3~7.6) are commonly used for

predicting the bioavailability or lability of metals such as Cu, Zn and Cd, in soil13-20. CaCl2 extraction

(0.01M) predicts metal bioavailability by mimicking the chemistry of soil pore water and targets the

exchangeable metal ions in soil pore water which are ‘readily available’ to plants in soils14, 20, 21. DTPA

is a strong chelating agent that mimics the chelating effect of root exudates to enhance the nutrient

availability from soil for subsequent uptake 15. The DTPA extraction not only targets the free ions in

soil pore water, but also the carbonate-bound and the organic-bound fractions of metal in soil,

which could be ‘potentially available’ to plants14, 22. While these extraction methods for assessing the

lability of Cu in Cu salt (CuSO4 and Cu(NO3)2) amended soil or for metal contaminated soils are

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well-developed, there are only a few reports using such methods with Cu-based NP or other

metal/metal oxide nanoparticles in soil23-25.

Recently, a few studies have used single time point CaCl2 extraction and DTPA extractions

to predict the lability of metal/metal oxide nanoparticles in soil. Judy et al.25 used CaCl2 and DTPA

extractions to estimate the bioavailability of ZnO-NP, TiO2-NP and Ag-NP in soil and concluded

that these extraction methods could not predict their bioavailability to plants (Medicago truncatula) in

Woburn sandy soil. Pradas del Real et al.23 used DTPA and CaCl2 extractions to assess the labile

pool of Ag in Ag NP amended soil, and concluded that the low extractability of Ag in soil was

consistent with the low bioavailability of Ag to plants (wheat and rape) in a loamy soil. Xu et al.24

used CaCl2, EDTA and DTPA extractions to estimate the bioavailability of CuO NP and TiO2 NP

to soil microbes and their community structures in a typical paddy soil. They observed that DTPA

and EDTA extractable Cu in CuO NP amended soil correlated well with microbial activity

(microbial biomass, soil enzyme activity, and total phospholipid fatty acids) in a CuO NP amended

soil. So far, results from studies on the use of chemical extraction methods to predict the

bioavailability of metals from nanoparticles in soils are contradictory and often inconclusive. One

reason for this may be the fact that these studies did not assess the rates of transformations of NP in

those soils and the corresponding effect on metal extractability. We hypothesize that aging time and

concentration will be important factors influencing these particles’ transformation and bioavailability

in soil, which may explain the absence of a correlation between extractability and bioavailability

using a single time-point extractions23-25.

The dissolution and transformation of some metal and metal oxide NP in soil have been

determined. The dissolution of copper oxide nanoparticles over time in three soils was reported by

McShane et al26. In their study, they measured an increase in free Cu2+ activity in soil pore water over

time and concluded that CuO NPs were dissolving. However, the rate of dissolution of CuO NP

was not modeled or reported. The present study extends this work by McShane et al. by measuring

pore water and SOM-associated Cu species using well-established extraction methods designed to

assess bioavailable fractions of Cu, by synchrotron X-ray analysis to confirm changes in copper

speciation, and by determining the effect of NP concentration on the dissolution behavior. The

transformations of metal and metal oxide nanoparticles in soil have been monitored using

synchrotron X-ray absorption spectroscopy (XAS)27-32 to measure changes in metal speciation over

time. Recently, Sekine et al.27 used XAS to monitor the change of speciation of Ag-NP, AgCl-NP

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and Ag2S-NP in soil over time. They observed that an increase in S-bound Ag species, including

Ag2S-NP, Ag-cysteine and Ag-cysteine, correlates with the decrease in labile Ag determined using

diffusive gradients thin films. However, the Ag NP transformation kinetics were not studied.

The dissolution of a number of metal and metal oxide nanoparticles in water has been

reported33-39. Most studies use empirical first-order dissolution models to describe their dissolution 35-

38, and evidence suggests that the measured dissolution rate constants are concentration dependent36.

However, the dissolution rate of metal and metal oxide nanoparticles in soils, where water content

and SOM can greatly affect the dissolution, is less well-understood. These rates are needed to

understand the dynamic nature of nanoparticulate metals relative to soluble metals added to soils

and to parameterize fate and transport models for engineered nanomaterials. 40

The objectives of the present study were to (a) compare the extractability of CuO NP with

the extractability of Cu(NO3)2 in soil, (b) quantify the extractability of CuO NP as a function of time

and nanoparticle concentration in a sandy (Lufa 2.1) soil (c) determine the fate processes influencing

the extractability of CuO NP in soil and (d) to model the dissolution kinetics of CuO NP in soil

from extraction experiments. We used 0.01 M CaCl2 and 0.005M DTPA (pH=7.6) extraction

methods to study the extractability of Cu(NO3)2 and CuO NP in aerated soils over a one-month

period at two different total added Cu concentrations (10 and 100 mg Cu kg-1 (mg/kg) dried soil).

Changes in speciation of Cu in soil were monitored using XAS to infer the dissolution of CuO NP.

2.2 Method and Materials

2.2.1 Chemicals

CuO NP (50 nm), DTPA, (>99% (titration)) and triethanolamine (TEA, ≥99.0% (GC)) were

purchased from Sigma-Aldrich. Cu(NO3)2 (>98% ACS grade), calcium chloride (≥99.0%, (ACS

grade)) and sodium bicarbonate (≥99.7%, (ACS grade)) were purchased from Fisher Scientific.

2.2.2Nanoparticle Characterization.

Primary particle size distribution of the CuO NP was characterized by transmission electron

microscopy (TEM, Hitachi H-9000 TEM microscope operating at 300 kV). The hydrodynamic

diameter and zeta potential of CuO NP in suspension (80 mg/kg as Cu in 5mM pH=7 NaHCO3

buffer) were determined by dynamic light scattering (Zetasizer Nano, Malvern). The isoelectric

points of 80mM CuO NP in 5mM NaHCO3 buffer and in 5mM NaNO3 were calculated from

- 14 -

measurements of the zeta potential of the particles in suspension over a range of pH. The crystal

structure of CuO NP was determined by X-ray powder diffraction (XRD, Panalytical X’Pert Pro

MPD X-Ray Diffractometer).

2.2.3 Soils and Characterization of Soil Properties. Standard soil (2.1-sandy soil) was purchased from Lufa, Germany. The standard soil (Lufa 2.1) was

used because it is commonly used in bioavailability studies and therefore can enhance comparison

from different studies. Lufa 2.1 soil also contains very little extractable Cu and total Cu (as discussed

later in ‘Soil and nanoparticle characterization’ section), making background interference minimal.

Lufa soil was air dried and sieved < 2mm before shipping. The soil was further air-dried for 12

hours before all experiments. Soil pH was determined according to the standard procedures

recommended by the USDA41. Specifically, 5 g of air-dried soil was mixed by hand for 10s with 5ml

of deionized water. The pH of the solution was measured after allowing the mixture to settle for 10

minutes. To determine the soil moisture content, 2 g of the air dried soil were dried in an oven at

105 ºC for 24 h 42. The moisture content was then determined gravimetrically. Soil field moisture

capacity was determined using a modified cylinder method in which air-dried soil was added to a

15ml-graduated cylinder. Deionized water was then added into the cylinder to wet the top 2 cm of

soil. After 24h, the wetting front in the soil moved downward. After removing the top 2 cm of soil,

the moisture content of soil above the wetting front (which was assumed to be at soil’s field

capacity) was determined.

2.2.4 Soil amendment and incubation. Two doses of CuO NP and Cu(NO3)2 were used in our study: 10mg Cu/kg dry soil for the low dose

amendment, and 100 Cu mg/kg dry soil for the high dose amendment. These two doses were

selected to investigate the influence of concentration on extractability of CuO NP in soil. While the

low does is more realistic, the high dose provided sufficient Cu concentration for XAS study. Soils

were amended with CuO NPs or Cu(NO3)2. All amended soil samples were incubated in 50ml

centrifuge tubes under aerobic conditions between 0 and 31 days before being extracted and

digested. Holes were made in the caps of centrifuge tubes for air exchange. This aerobic condition

was chosen to prevent CuO NP from being reduced to Cu(I) species or Cu(0). Cu speciation

analysis from XAS confirmed the absence of significant amounts of reduced Cu species in the

samples (<a few wt%) by. The experiments described in the following chapters of this thesis applied

the same incubation conditions for ENM soil amendments. Additional details of the amendment

procedure can be found in Appendix 1.

- 15 -

2.2.5 Total Metal Concentration. Soil total metal concentration was determined using acid digestion according to USEPA Method

3050B (1996). According to the procedure, 1g of air-dried soil was digested with concentrated nitric

acid and 30% hydrogen peroxide at 95 ºC using a hot plate. After digestion, the samples were

centrifuged at 3000 rpm for 10 min, followed by filtration using 0.45um filter to remove fine

particles in the supernatant. The filtered supernatant was diluted with Milli-Q water and acidified

with 20% HNO3 (final HNO3 concentration was 2%) for analysis by ICP-MS (Agilent 7700x). The

instrument was calibrated with a mixed calibration standard (purchased from Agilent Technologies)

every time before measurement. The calibration ranges used for different samples can be found in

table A1-1 (Appendix 1).

2.2.6 Extractions to assess the labile Cu in soil samples. After different incubation periods, 2.0 g of air-dried soils or 2.3 g of wet soils were extracted with

two standard extractants: The first one (termed DTPA) uses a 4 mL mixture of 0.01M CaCl2,

0.005M DTPA and 0.1M triethanolamine (TEA) (pH=7.6). The second one (termed CaCl2) (pH=5)

uses 20 mL of 0.01M CaCl2. All extractions were done using a reciprocal shaker at 180 rpm for 2

hours. Sample bottles were laid horizontally in the shaker. Both wet soil and air dried soil were used

to study the effect of air drying. After extraction, all samples were centrifuged at 3000 rpm for 10

min, and the supernatants were filtered with using a 0.2 um PTFE filter. In order to monitor the

impact of CuO NP suspension or Cu(NO3)2 solution on pH of soil, the pH of CaCl2 extracts for air-

dried amended soil and a unamended soil (no nanoparticle or Cu(NO3)2 added) were also measured

to estimate the soil pore water pH. The samples collected were further filtered with a 3kda filter to

separate the dissolved and nanoparticulate fraction of Cu in extracts. All samples were acidified with

20% HNO3 (final HNO3 concentration was 2%) and Milli-Q-water and analyzed by ICP-MS. Due to

the large difference between Cu concentrations from CaCl2 extracts and Cu concentration from

DTPA extracts, different calibration ranges were used. The different calibration ranges used for

different samples can be found in Table A1-1 in Appendix 1.

2.2.7 Determination of Cu speciation in soils. Cu speciation in soils (Lufa 2.1) on 1, 4, 7 and 19 days after amendment was analyzed by Cu K-edge

XAS at the Stanford Synchrotron Radiation Lightsource (SSRL) on Beamline 11-2. Spectra for both

100mg/kg and 10mg/kg amended soils were collected. However, the signal-to-noise ratio for the

10mg/kg amended soils was too poor for adequate speciation. Specifically, samples were lyophilized,

- 16 -

ground with a mortar and pestle to achieve uniformity, pressed into pellets, and placed between

Kapton tape. A double crystal Si (220) monochromator was calibrated by setting the first inflection

of the K-edge of a metallic Cu foil to 8979 eV. Harmonic rejection was achieved by detuning the

monochromator crystal by 25%. Spectra of soil samples were recorded in fluorescence mode at

room temperature using a 100-element germanium detector. The scans were averaged, energy

corrected using a metallic Cu foil standard, deadtime-corrected, background subtracted with E0

defined at 8988 eV, and de-glitched using SIXPack data analysis software43. Spectra were analyzed by

linear combination fitting (LCF) using the following reference spectra: CuO NP, metallic Cu,

CuSO4, Cu(NO3)2, CuPO4, Cu-cysteine, Cu2S (chalcocite mineral sample) ,CuS (covellite mineral

sample), Cu- iron oxide, Cu+ sorbed to humic acid (Cu(I)-HA) and Cu2+ sorbed to humic acid

(Cu(II)-HA). Inclusion of a reference spectrum into the combination fit required at least a 10%

decrease in the R-value, indicating a significant change to the quality of the fit.

2.2.8 Dissolution kinetics. For both extraction methods, the extractable Cu (either in pore water (CaCl2), or pore water plus soil

bound Cu (DTPA)) is assumed to increase proportionally as the CuO NPs dissolve.

The increase in the extractability of Cu over time is modeled using equation 2-1,

𝑑𝑑𝑑𝑑𝑑𝑑𝑑𝑑

= 𝑘𝑘�𝐸𝐸𝑓𝑓𝑓𝑓𝑓𝑓𝑓𝑓𝑓𝑓 − 𝐸𝐸� (2-1)

where E is the concentration of extractable Cu at time t, k is the empirical 1st order extraction rate

constant, and Efinal is the concentration of extractable Cu at the end of experiment. If the dissolution

of the CuO NP is the rate limiting step, i.e. the Cu-soil organic matter interaction is much faster

than the dissolution of CuO NP in soil, then the measured extraction rate constants from both

extractions should be similar, and equal to the CuO NP dissolution rate constant.

2.3 Results and Discussion 2.3.1 Soil and nanoparticle characterization. Lufa 2.1 soil is a sandy soil, containing 3 wt% clay, 11 wt% silt and 86 wt% sand (as provided by

Lufa). It has low organic matter content (organic carbon content is 0.7 wt% as provided by Lufa).

After air-drying, Lufa soil had 1.2 wt% moisture content. The soil pH was 5.6 and the field capacity

- 17 -

was 16 wt%. The total Cu concentration of the unamended soil was 2.95±0.11mg/kg. Total Cu

concentration measured in each of the amended soils is presented in Table A1-2 (Appendix). The

DTPA extractable Cu in unamended soils ranged from 0.37 to 0.53 mg/kg dried soil while the CaCl2

extractable Cu in unamended soils ranged from 0.005 to 0.024mg/kg (Table A1-3 in Appendix).

The primary particle size of CuO NP (measured from TEM) was 38nm (s.d. =14nm, 278 particles

were counted). The hydrodynamic diameter and zeta potential of 80mg/kg CuO NP in pH=7, 5mM

NaHCO3 buffer were 557nm (s.d. =56nm, 3 replicates, polydispersity index<0.297) and -16.1mv

(s.d. =0.8mV, 3 replicates), respectively. TEM images of the particles, along with the number of the

primary particles used to determine the size distribution are provided in the supporting information

(Figure S1-1, Appendix 1). Hydrodynamic size distribution (intensity averaged and number

averaged) and autocorrelation functions of CuO NP are also provided in supporting information

(Figure S1-2, Appendix 1). The isoelectric point of CuO NP shifted from pH=5.8 (in 5mM

NaHCO3 buffer) to pH=8.8 (in 5mM NaNO3), indicating a specific interaction between carbonate

and/or bicarbonate with the CuO NPs. The zeta potential of CuO NP measured over a range of pH

in NaHCO3 buffer and NaNO3 is provided in the supporting information (Figure A1-3 in Appendix

1). XRD results (Figure A1-4 in Appendix 1) indicate that the CuO NP we used is tenorite.

2.3.2 The change of soil pH after amendment. The pH of amended soils and unamended soils were stable over time. The pH of soil pore water

(measured in CaCl2 extracts) ranged from 4.9 to 5.1, except for the high dose CuO NP amended

soil, whose pH ranged from 5.2 to 5.5. The relatively higher pH in the high dose CuO NP amended

soils may be due to the acid-promoted dissolution of the CuO NP (eqn. 2-2 and 2-3). This

dissolution is described in more detail later in the manuscript.

CuO(s) + H2O(l) ↔ Cu(OH)2 (s) (2-2)

Cu(OH)2 (s)+2H+(aq) ↔ Cu2+

(aq) +2H2O(l) (2-3)

In our system, the pH rose from 5.1 to 5.4. This increase is less than expected for consumption of

40 mg Cu/kg soil of according to eqn 2-3, suggesting that the soil’s buffering capacity limited the

- 18 -

increase in soil pH44. The pH measured in the CaCl2 extracts can be found in supporting information

(Figure A1-5 in Appendix).

2.3.3 General trends in extractable Cu for CuO NP- and Cu(NO3)2-amended soil. The CaCl2 (0.01M) and DTPA extractable (0.005M, pH=7.6) Cu for CuO NP and Cu(NO3)2

amended soils over time are shown in Figure 2-1. Both the low dose (10 mg/kg) and high dose

(100mg/kg) scenarios are included for comparison. For both the low and high doses of added Cu,

there are clear differences in the trends of extractable Cu for CuO NP compared to Cu(NO3)2.

Initially, the extractable Cu for Cu(NO3)2 amended soils was higher than for the CuO NP amended

soil for both CaCl2 and DTPA extractions. Over time, the Cu(NO3)2 amended soils showed a

decrease in extractable Cu, whereas the CuO NP amended soils showed an increase in extractable

Cu with time. This is consistent with the findings of McShane et al.26 who found that ionic Cu in

pore water increased with time for CuO NP addition, but decreased with time for the Cu(NO3)2

addition. The trend of lower extractability of Cu in Cu(NO3)2 amended soil over time is well

documented; the lower extractability over time is a result of micro pore diffusion and the

complexation of ionic Cu by SOM10-12 as well as possible irreversible binding between ionic Cu and

SOM 45. The increasing extractability of Cu from CuO NP with time suggests that the NPs were

transforming to become more extractable. This is a result of dissolution (as suggested by McShane et

al.26), which was confirmed in this study with XAS, as discussed later in the paper. Note that the low

extractable Cu in CuO NP amended soil on day 0 (immediately after amendment) indicates that the

extractions did not induce nanoparticle dissolution.

For all extractions and time points, the DTPA extractable Cu was higher than CaCl2 extractable Cu

for both CuO NP amended soil and Cu(NO3)2 amended soil. This is consistent with former studies 17, 24, 25. This is because the CaCl2 extraction only extracts the dissolved metal and small particulate

metal in pore water, while DTPA extracts both the metal in pore water and the carbonate mineral-

bound and organic-bound metal15, 20-22. Importantly, this result suggests that dissolution of CuO NPs

is followed by an interaction between released copper and soil organic matter, which is known to

affect the amount of ionic Cu in soil. In the high dose addition, there was a relatively rapid change

occurring in the first ~7 days, followed by a period of slower change as the system approached an

apparent steady state with respect to CuO(s) dissolution. In the low dose addition, the extractability

of Cu for the CuO NP addition is the same as that for Cu(NO3)2 for t > 10 d (p > 0.05,

Kolmogorov-Smirnov test). This suggests that the Cu may be fully dissolved and “aging” similarly

- 19 -

to the Cu(NO3)2. However, the slight downward trend in extractability for t > 10d is not statistically

significant (P > 0.05, one-way ANOVA test).

Although extraction procedures generally used air dried soils21,46 , we used both the air dried soils

(after incubation) and wet soils for extractions to investigate the influence of air drying on

extractability of CuO NP in soil. Our results indicated that air drying has no significant effect

(P>0.05, Kolmogorov-Smirnov test) on extractability of Cu in both CuO NP amended and

Cu(NO3)2 amended soils. Thus, only the results from air dried soil is shown in Figure 2-1 for clarity.

Additional discussion on the effect of air drying can be found in the supporting information.

0 1 0 2 0 3 0 4 00

2

4

6

8

1 0

I n c u b a t i o n t i m e ( d a y s )

C u ( N O 3 ) 2

C u O N P

( a )

Ex

tra

cta

ble

Cu

(mg

/k

g d

rie

d s

oil

)

1 0 m g / k g , D T P A e x t r a c t i o n

0 1 0 2 0 3 0 4 00 . 0

0 . 1

0 . 2

0 . 3


C u ( N O 3 ) 2

C u O N P

( b )E

xtr

ac

tab

le C

u

(mg

/k

g d

rie

d s

oil

)

1 0 m g / k g , C a C l 2 e x t r a c t i o n

0 1 0 2 0 3 0 4 00

2 5

5 0

7 5

1 0 0


C u ( N O 3 ) 2

C u O N P

( c )

Ex

tra

cta

ble

Cu

(mg

/k

g d

rie

d s

oil

)

1 0 0 m g / k g D T P A e x t r a c t i o n

0 1 0 2 0 3 0 4 00

4

8

1 2


C u ( N O 3 ) 2

C u O N P

( d )

Ex

tra

cta

bl e

Cu

( mg

/k

g d

r ie

d s

oi l

)

1 0 0 m g / k g , C a C l 2 e x t r a c t i o n

Figure 2-1. Extractable Cu and in CuO NP and Cu(NO3)2 amended soils as a function of time and

the first order dissolution fit for CuO NP in soil: (a) DTPA extraction for 10 mg/kg amendment, (b)

CaCl2 extraction for 10 mg/kg amendment, (c) DTPA extraction for 100 mg/kg amendment and (d)

CaCl2 extraction for 100 mg/kg amendment. Error bars indicate ± 1 standard error. Dashed lines

indicate model fits using equation 1. For the low dose amendment, because CuO NPs were fully

dissolved after the 7-day sampling time, we modeled only the first 7 days. represents extractable

- 20 -

Cu in CuO NP amended soils air dried after incubation and represents extractable Cu in

Cu(NO3)2 amended soils air dried after incubation.

Figure 2-2. Fraction of small particles and dissolved ions (those passing 3kDa filter) in (a) DTPA

extracts and (b) CaCl2 extracts. D1, D2, D31 stand for 1 day, 2 days and 31 days after dosing. Error

bars indicate ± 1 standard error.

2.3.4 Fractions of dissolved Cu and particulate Cu in extracts. Bioavailability of Cu depends on its speciation, e.g. free ions, complexed ions and particulate

species47. We used filtration (first a 0.2-micron filter followed with a 3kDa filter) to distinguish

between dissolved and particulate species of Cu in each of the extracts. Figure 2-2 shows the

fraction of Cu that passes the 3 kDa filter (considered dissolved) in CaCl2 and DTPA extracts. For

DTPA extraction, nearly all extractable Cu (from 90% to 100%) was dissolved. This is because most

Cu in DTPA extracts bound with the chelating agent (DTPA) and the Cu-DTPA complex can pass

through the 3kDa filter. In contrast, filtration of the CaCl2 extract indicated the presence of Cu-

containing particles compared to the DTPA extracts (P<0.05, Kolmogorov-Smirnov test). These

small particles may include Cu2+ ion complexed with SOM or potentially small CuO NP (in CaCl2

extracts for CuO NP amended soil). The species of Cu in CaCl2 extracts was not analyzed in this

study. However, there were no effects of concentration, type of Cu added, or time on the amount

of dissolved vs. particulate Cu (p>0.05, one way ANOVA test).

- 21 -

2.3.5 Effect of CuO NP concentration on its extractability in soil. The concentration of added Cu influences the extraction behavior for CuO NP compared to

Cu(NO3)2. For the low Cu dose, the extractability of Cu in CuO NP amended soil was the same as

for the Cu(NO3)2 amended soil after ~10 days. No statistically significant difference (p>0.05,

Kolmogorov-Smirnov test) is found for extractable Cu for both CaCl2 extractions and DTPA

extractions between CuO NP amended soil and Cu(NO3)2 amended soil on day 13, 19 and 31,

suggesting that the CuO NP were fully dissolved before 13 days in soil at the lower dose. The

behavior was quite different at the high Cu dose. For the high dose of added Cu, extractable Cu in

Cu(NO3)2 amended soils was always higher than the extractable Cu in CuO NP amended soil. The

extractability of Cu from the CuO NP amended soil increased over the entire 31day period,

suggesting that CuO NP was dissolving over 31 days, but the dissolution of CuO NP in soil was not

complete. One possible explanation on the persistence of CuO NP and the slower dissolution rate

after ~7 days in the high dose soil (100mg/kg Cu) is that the free Cu2+ in soil pore water approached

saturation with respect to CuO(s). Conversely, the lower dose system (10 mg/kg) was not

oversaturated with respect to the CuO(s) phase. While CaCl2 extraction is a well-established method

to assess the pore water concentration of dissolved Cu, the potential for artefact during the

extraction and uncertainty in the complexation constants for Cu and the NOM in our system

prevents an accurate determination of the degree of saturation in the pore water. .

2.3.6 Dissolution rate of CuO NP in soil. For the high dose of CuO NP (100 mg/kg), the first-order extraction model describes the change of

extractable Cu over time well (R2>0.995) (dashed lines in Figure 2-1). However, we should note that

Cu2+ ions dissolved from CuO NP can become irreversibly bound with soil organic matter, making

it unextractable by DTPA, as indicated in former sections. This irreversible interaction is about 20%

for our soils, and has a minimal effect on the calculated dissolution rate constant. This is in part

because it is a small fraction of the total, and in part because the time scale for partitioning into this

irreversible fraction is short, i.e. less than 1d compared to the dissolution processes being

investigated, i.e. many days to weeks. For CaCl2 extraction, the fraction of extractable ionic Cu was

significantly less, with only 2% to 10% of the ionic Cu being extractable because it targeted only Cu

in soil pore water. Despite the differences in the extractable amount of Cu, the modeled dissolution

rate constants for DTPA extractable Cu and CaCl2 extractable Cu are similar (Table 2-1). This

indicates that the extractable amount of Cu by either the DTPA or CaCl2 extraction can be used to

monitor the CuO NP dissolution in the soils. This is a natural consequence of a first-order

- 22 -

dissolution process, which scale with the ratio of the final and initial concentration (C/Co) so any

process that reduced C and Co by the same constant fraction will not affect the calculated rate.

Moreover, it suggests that Cu2+ binding to SOM is rapid enough, such that dissolution of the CuO

NP is the rate-limiting process controlling both DTPA extractable Cu and CaCl2 extractable Cu in

soil.

2.3.7 Effect of aging on speciation of Cu in Cu(NO3)2 and CuO NP amended soil. Speciation of Cu in the 100mg/kg CuO NP and 100mg/kg Cu(NO3)2 amended soils were

determined at selected time points using XANES (Figure 2-3). Details regarding the spectra for

model compounds and fitting result can be found in the supporting information (Figure A1-7 and

Table A1-4 in Appendix 1). The speciation of Cu in Cu(NO3)2-amended soils can be adequately

modeled using only the Cu(II)-HA model compound, indicating that the Cu has predominantly Cu-

O character, i.e. associated with humic acids or potentially (but less likely) with clay or metal oxide

surfaces of the solids. This is consistent with prior speciation studies indicating that the main species

of Cu in soil is Cu(II)-HA using experimental approachs48-49 and with results of equilibrium

partitioning modeling50. This also suggests that Lufa 2.1 soil has the capacity to sorb up to

100mg/kg of added Cu, because our data showed that all Cu in the 100mg/kg Cu(NO3)2 amended

soil was Cu(II)-HA. In contrast, the Cu speciation in CuO NP amended soil required both Cu(II)-

HA and CuO NP model compounds. In the high dose CuO NP amended soil, linear combination

fitting indicates that the presence of CuO decreases over time, with a subsequent increase in the

Cu(II)-HA. This suggests that the CuO NPs were dissolving relatively fast in the first 7 days and

then more slowly after that as the pore water becomes saturated with respect to CuO(s). The rapid

dissolution in the first 7 days in consistent with the DTPA and CaCl2 extractability data, which

increased most rapidly in the first 7 days, followed by a slower increase. The dissolution of CuO NP

slowed down after 7 days even though the soil has not reached its capacity to adsorb Cu, which

confirms our former assumption that dissolution of CuO NP is the limiting factor controlling the

extractability of Cu from soil. For both the CuO NP and Cu(NO3)2 treatments, XANES analysis

showed no indication of Cu reduction in the soil, confirming our assumption that the experimental

condition was aerobic. Note that we also analyzed the 10mg/kg soils and the unamended soil

samples, but the signal-to-noise ratio was too poor for adequate speciation.

Table 2-1. Modeled first-order dissolution parameters for CuO NP amended soil.

- 23 -

Extraction type k (day-1) 95% confidence intervals

for k (day-1)

Half-life

(days)

E0a

(mg/kg)

Efinal

(mg/kg)

R2

High dose amendment

DTPA extraction

(dry soil)

0.15 0.11-0.19 4.6 3.35 37.4 0.995

CaCl2 extraction

(dry soil)

0.13 0.12-0.18 5.2 0.05 1.0 0.998

Low dose amendment

DTPA extraction

(dry soil)

0.16 0.06-0.25 4.5 0.36 6.71 0.936

CaCl2 extraction

(dry soil)

0.11 0.07-0.14 6.6 0.03 0.16 0.975

a: E0= initial extractable Cu at day 0 (intercept at y axis)

Figure 2-3. Change of Cu speciation in amended soils as inferred by XANES: in (a) Cu(NO3)2

amended soil and (b) CuO NP amended soil dosed at 100 mg/kg total Cu. The red dash lines are

- 24 -

fitted data while the black lines are experimental data. Model compounds used for the fits are below

the experimental spectra. The pie charts represent linear combination fits of the various model

compounds.

2.4 Environmental Implications The extractability of Cu from CuO NP-amended soils is different from that in soils dosed with Cu

ions as Cu(NO3)2, suggesting that the lability of CuO NP may be different from the lability of the

highly soluble Cu salts used as pesticides in soils. CuO NP was much less labile than Cu(NO3)2 in

soil immediately after they were added to the soil, but its lability increased over time. The differences

in lability between CuO NP and Cu(NO3)2 became negligible at low Cu doses (10 mg/kg) after

about 7 days, but differences in lability remained over 31 days for the high dose. The increase of the

labile pool of CuO NP over time was a result of their slow dissolution. Thus, our research shows

that dissolution is an important process controlling the extractability of CuO NP in soil, but the

dissolution rate and CuO NP persistence will be concentration dependent. Moreover, the aging time

in soil must be considered when assessing the lability or bioavailability of CuO NP in soils as was

also previously suggested by Sekine et al for Ag NP, Ag2S NP and AgCl NP and McShane et al. for

CuO NP (dosed at 500 mg/kg)26, 27, along with the total applied dose. If toxicity is purely the result

of the release of copper ion, the regulatory limit for applying nano CuO in agriculture could be

adjusted to consider its “slow release” behavior and concentration-dependent persistence. Because

of the relatively slow dissolution behavior of CuO NP, the regulatory limit for CuO NP could be

higher than that set for Cu salts. This is especially true if, with some additional surface modification,

the dissolution rate of Cu-based nanoparticles could be further reduced. Compared with a direct

spray application of Cu salt, a slow sustained release of ions from CuO NP may have lower

environmental impact to groundwater and rivers because particles have lower leachability and

mobility. On the other hand, if CuO NPs exhibit nanoparticle specific toxicity51,52, for higher doses

where CuO NPs persist, regulations will need to consider this persistence if CuO NPs show greater

toxicity than the Cu salts. Overall, the regulation of nano enhanced particles might be better based

on their dissolution rate at the applied dose, which could be easily determined with the methods

used in this study.

This work advances our understanding of the fate of CuO NP in several important ways. First, we

found CuO NP dissolution is the rate limiting step in controlling the increase of CaCl2 extractable

Cu and DTPA extractable Cu in CuO NP amended soil, indicating the dissolution process of CuO

- 25 -

NP in soil is much slower than the Cu-SOM interaction. Thus, we can monitor the dissolution of

CuO NP in soil from either the increase in dissolved Cu in soil pore water (as indicated by CaCl2

extraction) or increase in extractable Cu by DTPA extraction (dissolved Cu plus Cu bound to SOM

and carbonates). While McShane et al.26 suggested that soil pH is an important factor controlling the

dissolution of CuO NP in soil, we also suggest that the amount of SOM in soil may be as or more

important because it provided the sink for the released Cu in the soils used here. Secondly, our

research indicates that the concentration of soluble nanoparticles added to the soils can affect

temporal changes in Cu speciation, which in turn can affect the interpretation of exposure or toxicity

testing. At a low dose (10 mg/kg dried soil), CuO NPs became fully dissolved within 10 days. Thus,

at low doses, exposures to nanoparticles after ~10 days are not occurring and exposures and toxicity

testing would be expected to be consistent with a dissolved Cu species. Moreover, the Cu species

present was similar to Cu(II)-HA found in the natural soil so responses to CuO NP amended soils at

these low doses would likely be similar to exposures to native soils with the same Cu concentration.

However, using a higher CuO NP dose (100 mg/kg dried soil), about 40% of CuO NPs remained

undissolved after 31 days, potentially because the dissolution was limited by the solubility with

respect to CuO(s). In experiments using this high concentration, exposures and effects may be a

result of interactions with CuO NP and therefore different than for added ions or native soils.

Our research suggests that a single time point extraction after dosing soil may not be adequate for

predicting bioavailability unless that extraction is made at the same time as the end point of interest

(e.g. plant uptake). Rather, a time series of extractions after dosing may be more appropriate for

predicting the bioavailability of metal/metal oxide nanoparticles in soil. The time series of

extractions used here could be developed as a functional assay for studying the dissolution kinetics

of metal/metal oxide nanoparticles in soil. The functional assay approach has recently been

proposed as a means to empirically predict nanomaterial behaviors in complex media53. The method

that we developed is simple, and highly reproducible among the three replicates in our experiments.

The dissolution rate constant could be used for nanomaterial risk forecasting in soil system, as

suggested by Hendren et al53. Further studies need to confirm this method using different

metal/metal oxide nanoparticles in different soil systems. For example, several well-known

limitations of soil extractions methods, e.g. dilution effects, and the presence of an “irreversibly

bound” fraction of metal, exist. In the current study, the irreversibly bound fraction was relatively

low (<20%) and was achieved quickly such that is remained constant during the extraction process.

- 26 -

This enabled calculation of a first-order dissolution rate constant because it is independent of the

extractable amount as long as the extractable percentage is not changing over time. For soils where

the irreversibly bound metal fraction is changing at time scales similar to the NP dissolution, this will

complicate the analysis. However, a time-series extraction using the ionic salt can be used to control

for this “aging” effect. Here, we used XANES analysis to monitor Cu speciation. In our paper, we

determined that the two extractions that we used did not significantly influence the dissolution

process (there was low Cu extraction at Day 0), but this is not necessary true for all nanoparticles. A

preliminary experiment is required to prove that the extraction procedures do not induce significant

particle transformation. Regardless of these limitations, the excellent correlation between

extractability and XANES analysis showed that soil extraction methods are indeed good proxies for

Cu dissolution studies in aerobic soil.

Importantly, for CuO NP, we found that dissolution is the main processes controlling its lability in

aerobic soil. However, for other particles such as metallic Cu-NP, Ag-NP, or other reactive or

redox-sensitive nanomaterials, or different soil conditions (e.g. anaerobic) different processes may

also affect bioavailability. For example, sulfidation of Ag NP and Cu NP have been shown to affect

its properties, fate in soil54, and toxicity55-57. Oxidation may also be an important determinant of

lability in soil28.Future research is needed to better relate these different transformation processes

with nanoparticle lability or bioavailability.

2.5 References of Chapter 2 (1) Tegenaw, A.; Tolaymat, T.; Al-Abed, S.; El Badawy, A.; Luxton, T.; Sorial, G.; Genaidy, A.,

Characterization and potential environmental implications of select Cu-based fungicides and bactericides employed in U.S. markets. Environ. Sci. Technol. 2015, 49, (3), 1294-302.

(2) Giannousi, K.; Avramidis, I.; Dendrinou-Samara, C., Synthesis, characterization and evaluation of copper based nanoparticles as agrochemicals against Phytophthora infestans. RSC Adv. 2013, 3, (44), 21743-21752.

(3) Liu, R.; Lal, R., Potentials of engineered nanoparticles as fertilizers for increasing agronomic productions. Sci. Total Environ. 2015, 514, 131-139.

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(5) Giller, K. E.; Witter, E.; Mcgrath, S. P., Toxicity of heavy metals to microorganisms and microbial processes in agricultural soils: a review. Soil Biol. Biochem. 1998, 30, (10), 1389-1414.

(6) Nagajyoti, P.; Lee, K.; Sreekanth, T., Heavy metals, occurrence and toxicity for plants: a review. Environ. Chem. Lett. 2010, 8, (3), 199-216.

(7) Posthuma, L.; Van Straalen, N. M., Heavy-metal adaptation in terrestrial invertebrates: a review of occurrence, genetics, physiology and ecological consequences. Comp. Biochem. Physiol. C: Pharmacol. Toxicol. 1993, 106, (1), 11-38.

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(8) Rodrigues, S.; Trindade, T.; Duarte, A.; Pereira, E.; Koopmans, G.; Römkens, P., A framework to measure the availability of engineered nanoparticles in soils: Trends in soil tests and analytical tools. TrAC, Trends Anal. Chem.2016, 75, 129-140.

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(11) Lu, A.; Zhang, S.; Qin, X.; Wu, W.; Liu, H., Aging effect on the mobility and bioavailability of copper in soil. J. Environ. Sci. 2009, 21, (2), 173-178.

(12) Ma, Y.; Lombi, E.; Oliver, I. W.; Nolan, A. L.; McLaughlin, M. J., Long-term aging of copper added to soils. Environ. Sci. Technol. 2006, 40, (20), 6310-6317.

(13) Rao, C.; Sahuquillo, A.; Sanchez, J. L., A review of the different methods applied in environmental geochemistry for single and sequential extraction of trace elements in soils and related materials. Water Air Soil Pollut. 2008, 189, (1-4), 291-333.

(14) Peijnenburg, W. J.; Zablotskaja, M.; Vijver, M. G., Monitoring metals in terrestrial environments within a bioavailability framework and a focus on soil extraction. Ecotoxicol. Environ. Saf. 2007, 67, (2), 163-79.

(15) Menzies, N. W.; Donn, M. J.; Kopittke, P. M., Evaluation of extractants for estimation of the phytoavailable trace metals in soils. Environ. Pollut. 2007, 145, (1), 121-130.

(16) Feng, M. H.; Shan, X. Q.; Zhang, S.; Wen, B., A comparison of the rhizosphere-based method with DTPA, EDTA, CaCl2, and NaNO3 extraction methods for prediction of bioavailability of metals in soil to barley. Environ. Pollut. 2005, 137, (2), 231-40.

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(19) Brun, L.; Maillet, J.; Richarte, J.; Herrmann, P.; Remy, J., Relationships between extractable copper, soil properties and copper uptake by wild plants in vineyard soils. Environ. Pollut. 1998, 102, (2), 151-161.

(20) Peijnenburg, W.; Jager, T., Monitoring approaches to assess bioaccessibility and bioavailability of metals: matrix issues. Ecotoxicol. Environ. Saf. 2003, 56, (1), 63-77.

(21) Houba, V. J. G.; Temminghoff, E. J. M.; Gaikhorst, G. A.; van Vark, W., Soil analysis procedures using 0.01Mcalcium chloride as extraction reagent. Commun. Soil Sci. Plant Anal. 2000, 31, (9-10), 1299-1396.

(22) Sahuquillo, A.; Rigol, A.; Rauret, G., Overview of the use of leaching/extraction tests for risk assessment of trace metals in contaminated soils and sediments. TrAC, Trends Anal. Chem.2003, 22, (3), 152-159.

(23) Pradas del Real, A. E.; Castillo-Michel, H. A.; Kaegi, R.; Sinnet, B.; Magnin, V.; Findling, N.; Villanova, J.; Carriere, M.; Santaella, C.; Fernandez-Martinez, A., Fate of Ag-NPs in sewage sludge after application on agricultural soils. Environ. Sci. Technol. 2016, 50, (4), 1759-1768.

(24) Xu, C.; Peng, C.; Sun, L.; Zhang, S.; Huang, H.; Chen, Y.; Shi, J., Distinctive effects of TiO2 and CuO nanoparticles on soil microbes and their community structures in flooded paddy soil. Soil Biol. Biochem. 2015, 86, 24-33.

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nodulation, shift microbial community composition, and result in increased metal uptake relative to bulk/dissolved metals. Environ. Sci. Technol. 2015, 49, (14), 8751-8758.

(26) McShane, H. V.; Sunahara, G. I.; Whalen, J. K.; Hendershot, W. H., Differences in soil solution chemistry between soils amended with nanosized CuO or Cu reference materials: implications for nanotoxicity tests. Environ. Sci. Technol. 2014, 48, (14), 8135-8142.

(27) Sekine, R.; Brunetti, G.; Donner, E.; Khaksar, M.; Vasilev, K.; Jamting, A. K.; Scheckel, K. G.; Kappen, P.; Zhang, H.; Lombi, E., Speciation and Lability of Ag-, AgCl-, and AgS-Nanoparticles in Soil Determined by X-ray Absorption Spectroscopy and Diffusive Gradients in Thin Films. Environ. Sci. Technol. 2014, 49, (2), 897-905.

(28) Gomes, S. I.; Murphy, M.; Nielsen, M. T.; Kristiansen, S. M.; Amorim, M. J.; Scott-Fordsmand, J. J., Cu-nanoparticles ecotoxicity–Explored and explained? Chemosphere 2015, 139, 240-245.

(29) Collins, D.; Luxton, T.; Kumar, N.; Shah, S.; Walker, V. K.; Shah, V., Assessing the impact of copper and zinc oxide nanoparticles on soil: a field study. PLoS One 2012, 7, (8), e42663.

(30) Unrine, J. M.; Tsyusko, O. V.; Hunyadi, S. E.; Judy, J. D.; Bertsch, P. M., Effects of Particle Size on Chemical Speciation and Bioavailability of Copper to Earthworms (Eisenia fetida) Exposed to Copper Nanoparticles. J. Environ. Qual. 2010, 39, (6), 1942-1953.

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(34) 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, 50, (13), 6772-6781.

(35) Bian, S.-W.; Mudunkotuwa, I. A.; Rupasinghe, T.; Grassian, V. H., Aggregation and dissolution of 4 nm ZnO nanoparticles in aqueous environments: influence of pH, ionic strength, size, and adsorption of humic acid. Langmuir 2011, 27, (10), 6059-6068.

(36) Zhang, W.; Yao, Y.; Sullivan, N.; Chen, Y., Modeling the primary size effects of citrate-coated silver nanoparticles on their ion release kinetics. Environ. Sci. Technol. 2011, 45, (10), 4422-4428.

(37) Mudunkotuwa, I. A.; Rupasinghe, T.; Wu, C.-M.; Grassian, V. H., Dissolution of ZnO nanoparticles at circumneutral pH: a study of size effects in the presence and absence of citric acid. Langmuir 2011, 28, (1), 396-403.

(38) Peretyazhko, T. S.; Zhang, Q.; Colvin, V. L., Size-controlled dissolution of silver nanoparticles at neutral and acidic pH conditions: kinetics and size changes. Environ. Sci. Technol. 2014, 48, (20), 11954-11961.

(39) Jiang, C.; 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, (19), 11476-11484.

(40) Dale, A. L.; Casman, E. A.; Lowry, G. V.; Lead, J. R.; Viparelli, E.; Baalousha, M., Modeling nanomaterial environmental fate in aquatic systems. Environ. Sci. Technol. 2015, 49, (5), 2587-2593.

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(41) J. B. Peters, M. V. N., C. A. M. Laboski, pH and Lime Requirement. In Recommended Chemical Soil Test Procedures for the North Central Region, North Central Regional Research Publication No. 221 (Revised): 2012, pp 4.1-4.7.

(42) Ghani, A.; Dexter, M.; Perrott, K., Hot-water extractable carbon in soils: a sensitive measurement for determining impacts of fertilisation, grazing and cultivation. Soil Biol. Biochem. 2003, 35, (9), 1231-1243.

(43) Webb, S., SIXpack: a graphical user interface for XAS analysis using IFEFFIT. Phys. Scr. 2005, (T115), 1011.

(44) Eckert, D.; Sims, J. T., Recommended soil pH and lime requirement tests. Recommended soil testing procedures for the northeastern United States. Northeast Regional Bulletin 1995, 493, 11-16.

(45) Mao, L.; Young, S. D.; Bailey, E. H. Lability of copper bound to humic acid. Chemosphere 2015, 131, 201–208.

(46) Ure, A., Single extraction schemes for soil analysis and related applications. Sci. Total Environ. 1996, 178, (1), 3-10.

(47) Nolan, A. L.; Mclaughlin, M. J.; Mason, S. D., Chemical speciation of Zn, Cd, Cu, and Pb in pore waters of agricultural and contaminated soils using Donnan dialysis. Environ. Sci. Technol. 2003, 37, (1), 90-98.

(48) Jacobson, A. R.; Dousset, S.; Andreux, F.; Baveye, P. C., Electron microprobe and synchrotron X-ray fluorescence mapping of the heterogeneous distribution of copper in high-copper vineyard soils. Environ. Sci. Technol. 2007, 41, (18), 6343-6349.

(49) Strawn, D. G.; Baker, L. L., Speciation of Cu in a contaminated agricultural soil measured by XAFS, μ-XAFS, and μ-XRF. Environ. Sci. Technol. 2007, 42, (1), 37-42.

(50) Bonten, L. T. C.; Groenenberg, J. E.; Weng, L.; van Riemsdijk, W. H. Use of speciation and complexation models to estimate heavy metal sorption in soils. Geoderma 2008, 146 (1–2), 303–310.

(51) Yuan, J.; He, A.; Huang, S.; Hua, J.; Sheng, G. D. Internalization and Phytotoxic Effects of CuO Nanoparticles in Arabidopsis thaliana as Revealed by Fatty Acid Profiles. Environ. Sci. Technol. 2016, 50 (19), 10437–10447.

(52) Mashock, M. J.; Zanon, T.; Kappell, A. D.; Petrella, L. N.; Andersen, E. C.; Hristova, K. R. Copper Oxide Nanoparticles Impact Several Toxicological Endpoints and Cause Neurodegeneration in Caenorhabditis elegans. PLoS One 2016, 11 (12), e0167613.

(53) 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.

(54) Lombi, E.; Donner, E.; Taheri, S.; Tavakkoli, E.; Jämting, Å. K.; McClure, S.; Naidu, R.; Miller, B. W.; Scheckel, K. G.; Vasilev, K., Transformation of four silver/silver chloride nanoparticles during anaerobic treatment of wastewater and post-processing of sewage sludge. Environ. Pollut. 2013, 176, 193-197.

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(56) Starnes, D. L.; Unrine, J. M.; Starnes, C. P.; Collin, B. E.; Oostveen, E. K.; Ma, R.; Lowry, G. V.; Bertsch, P. M.; Tsyusko, O. V., Impact of sulfidation on the bioavailability and toxicity of silver nanoparticles to Caenorhabditis elegans. Environ. Pollut. 2015, 196, 239-246.

(57) Ma, R.; Stegemeier, J.; Levard, C.; Dale, J. G.; Noack, C. W.; Yang, T.; Brown, G. E.; Lowry, G. V., Sulfidation of copper oxide nanoparticles and properties of resulting copper sulfide. Environ. Sci. Nano. 2014, 1, (4), 347-357.

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CHAPTER 3: Quantify the effect of soil properties, including soil moisture content, organic carbon content and pH, on the dissolution

kinetics of CuO NP in soil. Abstract: The objectives of this research were to quantify the impact of organic matter content, soil pH and

moisture content on the dissolution rate and solubility of copper oxide nanoparticles (CuO NPs) in

an aerobic soil, and to develop an empirical model to predict the dissolution kinetics of CuO NPs in

an aerobic soil. CuO NPs were dosed into standard LUFA soils with various moisture content, pH

and organic carbon content. Chemical extractions were applied to measure the CuO NP dissolution

kinetics. Doubling the reactive organic carbon content in LUFA 2.1 soil increased the solubility of

CuO NP 2.7-fold but did not change the dissolution rate constant. Increasing the soil pH from 5.9

to 6.8 in LUFA 2.2 soil decreased the dissolution rate constant from 0.56 mol1/3·kg1/3·s-1 to 0.17

mol1/3·kg1/3·s-1 without changing the solubility of CuO NP in soil. For six soils, the solubility of CuO

NP correlated well with soil organic matter content (R2 = 0.89) independent of soil pH. In contrast,

the dissolution rate constant correlated with pH for pH<6.3 (R2 = 0.89). These relationships

predicted the solubility and dissolution rate constants of CuO NP in two test soils (pH=5 and

pH=7.6). Moisture content showed negligible impact on the dissolution kinetics of CuO NPs. Our

study suggests that soil pH and organic matter content affect the dissolution behavior of CuO NP in

soil in a predictable manner.

This work has been published in Environmental Science & Technology as ‘Effect of Soil Organic

Matter, Soil pH, and Moisture Content on Solubility and Dissolution Rate of CuO NPs in Soil’ , doi:

acs.est.8b07243

- 31 -

3.1 Introduction

Copper(Cu)-based nanoparticles (NPs) have been used in agriculture as fungicides and have

potential for use as trace element fertilizers1–5. Either through terrestrial application or foliar

application, these NPs will intentionally or inadvertently enter soils, where their fate depends

primarily on dissolution processes 6–8. Given that the release of Cu2+ ions is the primary mode of

toxicity of the copper oxide nanoparticles (CuO NPs)9,10, dissolution will affect both bioavailability

and toxicity of Cu-based NPs to plants and soil organisms11. Thus, dissolution rate and solubility are

properties of practical relevance12.

Solubility is the maximum mass of the CuO NPs that will dissolve and partition into different soil

compartments at equilibrium, and dissolution rate as the speed at which the ion release processes

occurs. These processes are influenced by environmental factors. Although the impact of matrices’

physical-chemical properties (e.g. concentration of organic matter (OM), pH, ionic strength) on the

dissolution rate of metal and metal (hydr)oxide nanoparticles in aquatic systems has been reported13–

16, such impacts in soils have not yet been systematically studied. Finally, while ionic Cu release from

Cu-based NP has been assessed in pore water7,17 and speciation in soil evaluated at a few given time

points6,8, abiotic factors influencing these dissolution behaviors have not been investigated yet,

resulting in the absence of adequate models to predict CuO NP dissolution kinetics in soils.

Measuring the dissolution of NPs in soil is not straightforward and studying their dissolution

kinetics is even more challenging. Recently, two methods have been developed to study the

dissolution of NPs in soil. The first method monitors the change in speciation of the metal in soil

over time using X-ray absorption spectroscopy (XAS)18,19. Transformation and dissolution is inferred

from an observed decrease in the fraction of the original NPs 6,8,18,20. Due to constraints on

synchrotron access and the relatively high cost of measuring each sample, such studies rarely provide

enough data to quantify the dissolution kinetics of NPs in soil, which requires multiple

measurements over time. The high cost also precludes XAS as a routine method for studying NP

dissolution processes.

Two soil extraction methods, developed originally to study the geochemical equilibrium distribution

of metals in soil and to predict their bioavailability to plants21, offer a less costly alternative. The first

method involves extracting pore water metals from soil using either dilute salt (e.g. 0.01M calcium

chloride, CaCl2 )22–25 or water7,26. It is used as a proxy for extracting soil pore water metal from

soil22,27,28. The extracted metal is considered to be ‘readily available’ to plants or ‘highly mobile Cu’.

- 32 -

The second extraction method is for labile metal and extracts ‘potentially available’ metals, i.e., those

that are reversibly bound to the soil solid matrix. Such extractions use chelating agents or dilute

strong acid (e.g., 0.005M Diethylenetriamine pentaacetate, DTPA, or 0.05M

Ethylenediaminetetraacetic acid, EDTA)24,25,29–31. The DTPA extraction method has been

demonstrated to extract most of the ionic Cu released from CuO NP6,10. The change over time in

extractability of metals in soil can be used as a proxy of metal NP dissolution. Both methods have

been assessed and can be used to monitor the dissolution processes of Cu based NPs in soil6,7.

Dissolution of metal and metal oxide NPs (e.g. CuO NPs) is most often an acid-promoted process.

Thus, the concentration of hydrogen ion (pH) plays a role in both the expression of the dissolution

rate law, and the equilibrium constant (Equation 3-1). OM, as well as other natural metal chelators,

(e.g., siderophores, or amino acids), can also affect dissolution by binding with metal ions released

by the NPs or by interacting with the particle (e.g. through coatings32) and affecting the total surface

available for dissolution. The dissolution of NPs also requires the interaction between NP and the

soil pore water. Given that moisture content in agricultural and natural soils varies over time and

space, it is also important to know how moisture content affects the dissolution of CuO NP in soil.

Redox potential is another influence to consider for the dissolution of CuO NP in soil. However,

for most agricultural soils, the plow layer, where agrochemicals are applied, is intended to be aerobic

(the redox potential is usually above 400mv due to the interaction with air33). At this redox

condition, reduction of Cu is not thermodynamically favored34. Thus, the influence of redox

potential on dissolution of CuO NP was not explicitly addressed in this work.

𝐶𝐶𝐶𝐶𝑂𝑂 𝑁𝑁𝑁𝑁𝑁𝑁(𝑁𝑁) + 2𝐻𝐻+(𝑎𝑎𝑎𝑎) ↔ 𝐶𝐶𝐶𝐶2+(𝑎𝑎𝑎𝑎) + 𝐻𝐻2𝑂𝑂 (3 − 1)

Recent efforts have been made to evaluate the environmental factors influencing dissolution kinetics

of CuO NP in soil. McShane et al. showed that the free Cu2+ concentration from CuO NPs

(measured using an ion selective electrode) in solution extracted with water from soil was affected by

soil pH, concluding that pH affected CuO NP dissolution in soil7. The correlation of free Cu2+ in

solution with dissolution is consistent with expectations based on CuO NP solubility in water.

However, previous studies have shown that pH also affects the partitioning of Cu2+ between pore

water and the soil solid surfaces. Higher pH results in more Cu2+ binding to soil organic matter

(SOM)29,35. Therefore, it is necessary to extract the dissolved Cu from the soils to ensure that all of

the dissolved Cu is accounted for in the measurement6. Another study used XAS to track the

- 33 -

changes in speciation of Cu in CuO-NP-amended soil. They also found that lower pH resulted in

higher dissolution in the short term (within 5 days) and that the SOM content slowed the dissolution

process in the short term8. The latter finding still requires more investigation since it contradicts the

findings from studies in water showing that SOM increased CuO NP dissolution13,15. SOM is an

important Cu sink/pool in soil36 because nitrogen and oxygen atoms in SOM can strongly bond

with Cu37, thus one would expect it should increase the solubility of CuO NP in soil. The studies

mentioned above also used a variety of soils to demonstrate the effect of soil pH and SOM on the

dissolution of CuO NP. This introduces potentially confounding variables as environmental factors,

e.g. soil texture and field capacity could also potentially affect the dissolution of CuO NP by

affecting CuO NP-soil aggregation and the distribution of CuO NP between soil pore water and soil

solids.

The objective of this study is to quantify the effect of pH, SOM content, and moisture content on

the dissolution rate and solubility of CuO NP in soil. We used several standard agricultural soils at

different pH and with different moisture and SOM content to investigate the influence of these soil

properties on CuO NP dissolution behavior. Dissolution models were then used to quantify the

effect of soil pH and SOM content on the dissolution kinetics of CuO NP in soil. Finally, the ability

of this model to predict the dissolution kinetics and solubility of CuO NP in soil was evaluated using

two test soils with different properties.


3.2.1 Chemicals

Calcium chloride (≥99.0%, ACS grade), calcium oxide (CaO), calcium carbonate (CaCO3) (99%+),

and hydrogen peroxide (30%, certified ACS) were purchased from Fisher Scientific. DTPA (>99%)

and triethanolamine (TEA, ≥99.0% (GC)) were purchased from Sigma-Aldrich. Trace metal grade

nitric acid (65%-70%) was purchased from VWR. Copper sulfate (CuSO4) was purchased from

Fisher Scientific . Lufa Standard soils (2.1, 2.2, 2.4 and 2.4) were purchased from Lufa Speyer,

Germany. A calcareous soil (pH 7.6) was collected in Arizona (termed Arizona soil) and used to test

the model’s ability to predict CuO NP dissolution behavior based on soil pH and SOM content.

Another more acidic soil (pH=5.0) was collected from a grassland in northwestern Portugal (termed

Portugal soil). Detailed properties of all the soils used can be found in appendix ( Table A2-1).

- 34 -

3.2.2Nanoparticle Characterization

CuO NPs (~40 nm primary particle size, zeta potential (ζ) = -16.1 mV ± 1.7mV at pH=7 in 5mM

NaNO3), were purchased from Sigma-Aldrich. The primary size of particles, zeta potential,

isoelectric point and hydrodynamic diameter have been characterized and reported in Chapter 26.

3.2.3 Soil amendment Soil pH, SOM content and moisture content, factors hypothesized to affect dissolution kinetics of

CuO NP in soil, were systematically varied in this study (the soil properties for all treatments can be

find in Table A2-2, Appendix 2). To investigate the effect of pH on the dissolution of CuO NP, a

mixture of CaO and CaCO3 powders were used to increase the soil pH from the original pH of 5 to

~7.5 for Lufa 2.1 soil (0.27g CaO, 0.68g CaCO3 in 270g of Lufa 2.1 soil), and from 5.9 to 6.8 for

Lufa 2.2 soil (0.27g CaCO3 in 270g of Lufa 2.2 soil)38. To investigate the influence of SOM on

dissolution of CuO NP with all other soil properties held constant, the soil total organic carbon

(TOC) content in Lufa 2.1 soil was increased from the original 0.7% to 0.9% by adding SOM

extracted from Lufa 2.1 soil. Note that generally the SOM content is ≈ 1.74 times the soil organic

carbon content, although this can vary between soil types39. SOM was extracted from Lufa 2.1 soil

following a procedure described by van Zomeren et al.40 Additional details on SOM extraction,

recovery, and preliminary characterization are provided in Appendix 2. Only about 23% of organic

carbon in Lufa 2.1 soil was extractable. This 23% is considered to be the ‘reactive organic carbon,’

the SOM fraction that usually controls the Cu sorption behavior. The remaining fraction was mostly

humic substances that have low affinity for metals41. In this study, 161mg extracted fulvic acid, FA,

and 368mg extracted humic acid, HA, was added to 90g Lufa 2.1 soil. In the original soil

(TOC=0.7%), the reactive carbon content was 0.16%. Thus, by adding 0.2% of reactive organic

carbon content in soil, the total reactive carbon in Lufa 2.1 soil was effectively doubled. (Note

carbon content in HA and FA are provided in Appendix 2.) CuO NPs and CuSO4 (control

treatment) were added to different soils to achieve final concentrations of 100 mg/kg, 250 mg/kg

and 500 mg/kg dry weight (dw) (as Cu). To investigate the influence of moisture content with all

other soil properties held constant, we used Lufa 2.2 standard soil at 21% and 10% moisture

content. The two moisture contents were selected because they span relevant moisture conditions,

on one end where the soil is as wet as it could be (field capacity) and the other as dry as it could

reasonably be (wilting point) for an agricultural soil. CuO NPs were also dosed into the Arizona soil

(500mg/kg Cu dw) and Portugal soil (500mg/kg Cu dw) to test our models’ ability to predict

solubility and dissolution rate of CuO NP in natural soils. The concentration of CuO used in each

- 35 -

treatment was selected based on the solubility of the CuO NPs in each soil determined in

preliminary studies (Appendix 2). Enough CuO NPs was added to each treatment to ensure that

some CuO NPs remained undissolved after 30d. Details on the treatment condition and Cu mass

balance are in appendix, Table SA-2. Note that during the 30d incubation period, all soils were

maintained under aerobic conditions (soils were incubated in centrifuge tubes with holes in the caps

allowing air exchange). It was verified in a previous study that these experimental conditions

precluded significant Cu reduction 42. For topsoil in agriculture (Eh>400mv)53,Cu(II) is the major Cu

valence state54. Thus Cu in or released from the CuO NPs are is assumed to remain in the Cu(II)

redox state.

3.2.4 Extraction procedure to measure the fraction of dissolved CuO NP and soil pH. The amount of CuO NP that had dissolved at each incubation time (days 0, 2, 4, 7, 14, 21, 30 after

amendment) and the corresponding soil pH at that time point, were measured using a previously

published extraction method6. Briefly, for each Cu treatment, 2.0 g of air-dried soils were extracted

with two standard extractants: (1) 4 mL of DTPA (0.05 M DTPA, 0.01M CaCl2 and 0.1M TEA at

pH 7.6) and (2) 20 mL of 0.01 M CaCl2 (pH =5). All extractions were done in a reciprocal shaker at

180 rpm for 2 hours. After extraction, samples were centrifuged and filtered with 0.45µm PTFE

filters. Then, the filtered samples were acidified and analyzed by ICP-MS (Agilent technologies

7700). The measurements were made right after each aging period. It should be noted that our

previous studies have demonstrated that such extractions did not induce any CuO NP dissolution6.

The pH of CaCl2 extracts for air-dried amended soils were measured as soil pH using a common

procedure43,44.

3.2.5 Determination of Cu speciation in soils

Cu speciation in soils after amendment was analyzed by Cu K-edge XAS at the Stanford

Synchrotron Radiation Lightsource (SSRL) on Beamline 11-2. Details on sample preparation and

measurements can be found in the appendix.

3.2.6 Dissolution models.

The model used for CuO NP dissolution in soil includes the following steps (Figure 3-1): (1): CuO

NP dissolves (reversibly, with rate constants kd and kr)), releasing free Cu ions into the soil pore

water. (2): Cu2+ attaches to different ligands (e.g. dissolved organic matter (DOM)) and soil surfaces

(e.g. clay, SOM) 45. The second step (Cu ion partitioning between soil pore water and soil solid

- 36 -

surfaces) has been investigated previously 29,46–48. The reversible dissolution of CuO NPs are of

primary interest to this study.

Figure 3-1. Schematic of CuO NP dissolution model. Where 𝒌𝒌𝒅𝒅 is the dissolution rate constant, 𝒌𝒌𝒓𝒓

is the local reverse reaction rate constant for Cu(II) ions precipitating back onto the CuO NP

surface (precipitation). Note that this reverse reaction must be occurring locally near the CuO NP

surfaces if the particles are in local equilibrium with the surrounding water. 𝑲𝑲𝒍𝒍𝒍𝒍𝒍𝒍𝒍𝒍𝒍𝒍𝒅𝒅 is the

partitioning constant between Cu associated with natural ligands (including both DOM and soil

surfaces, e.g. SOM, clay, iron oxides) and free Cu2+(aq). 𝒌𝒌𝒍𝒍𝒍𝒍𝒍𝒍𝒍𝒍𝒍𝒍 is the constant to account for

irreversible loss of Cu to the matrix over long time spans. It should be noted that only the CuO NP

dissolution parameters, highlighted in purple, are new additions to the well-known multi-surface

geochemical model47,49.

To model the dissolution kinetics, we define Cu2+Tot as the total concentration of Cu2+ being released

from CuO NP (free 𝐶𝐶𝐶𝐶2+ + 𝐶𝐶𝐶𝐶 𝑎𝑎𝑁𝑁𝑁𝑁𝑜𝑜𝑜𝑜𝑜𝑜𝑎𝑎𝑜𝑜𝑜𝑜𝑑𝑑 𝑤𝑤𝑜𝑜𝑜𝑜ℎ 𝑛𝑛𝑎𝑎𝑜𝑜𝐶𝐶𝑛𝑛𝑎𝑎𝑛𝑛 𝑛𝑛𝑜𝑜𝑙𝑙𝑎𝑎𝑛𝑛𝑑𝑑𝑁𝑁) , which can be extracted by

DTPA. If we assume that Cu2+Tot (t=0) = 0 and that [H+] remains constant during the dissolution

(implying a stable pH during the dissolution process due to the relatively high buffering capacity of

soil6), the rate law can be expressed by equation (3-2).

𝑑𝑑[𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑇𝑇𝑇𝑇,𝑇𝑇𝑑𝑑𝑑𝑑

= 𝑘𝑘𝑑𝑑([𝐶𝐶𝐶𝐶𝑂𝑂]𝑜𝑜 − [𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑜𝑜𝑑𝑑,𝑑𝑑)2/3 − 𝑘𝑘𝑟𝑟[𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑜𝑜𝑑𝑑,𝑑𝑑1

1+𝐾𝐾𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙([𝐶𝐶𝐶𝐶𝑂𝑂]𝑜𝑜 − [𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑜𝑜𝑑𝑑,𝑑𝑑)2/3

(3-2)

- 37 -

The derivation of Equation (3-2) can be found in Appendix 2. The key assumptions are:

1: The Cu2+ released by CuO NP is in equilibrium with respect to its partitioning to other soil

components, e.g. DOM and SOM. This equilibrium is fast compared to the rate of dissolution.

2: The solubility of CuO NP(s) is limited by the local dissolution/precipitation equilibrium.

The dissolution of CuO NP(s) in soils is not complete. A reverse reaction, which is a precipitation

process, must occur at the surface of CuO NP(s) to stop CuO NP from completely dissolving.

Dissolution stops when the dissolution rate near the CuO NP surface equals the reverse reaction

rate near the NP surface. The precipitation of Cu2+ preferentially happens near the surface of CuO

NP because of the localized higher Cu2+ concentration on the surface of the NP.

3: We assume that precipitation of Cu phases other than CuO is not significant.

This was corroborated with the facts that (a) ~80% of Cu was still extractable by DTPA in the Lufa

2.2 soil amended with a high concentration of CuSO4 (500 mg/kg), which did not form a solid

phase6; and (b) the Cu X-ray absorption near edge structure (XANES) spectra of Lufa 2.2 soil dosed

with 500mg/kg CuSO4 indicated that 99.6% of the Cu was present as Cu-NOM after 30 days

(appendix, Figure A2-1). It should be noted that the process of Cu2+ sorbed to the soil organic

matter (SOM) is not considered a ‘precipitation’ process, rather, it is a sorption process.

4: We assume the dissolution/precipitation of CuO NP are both surface-controlled process, e.g.

dissolution rate and the reverse reaction rate are both proportional to the total surface area of CuO

NP. Moreover, we assume that the CuO NPs are spherical and that their surface area changes

according to a 2/3 power law as has been previously described with the dissolution of spherical

ZnO NPs15.

At equilibrium, 𝑑𝑑[𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑇𝑇𝑇𝑇,𝑇𝑇𝑑𝑑𝑑𝑑

= 0 so the solubility of the CuO NPs in the soil, [𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑜𝑜𝑑𝑑,∞, is given

by Equation (3).

[𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑜𝑜𝑑𝑑,∞ =𝑘𝑘𝑑𝑑𝑘𝑘𝑟𝑟

(1 + 𝐾𝐾𝑓𝑓𝑓𝑓𝑙𝑙𝑓𝑓𝑓𝑓𝑑𝑑) (3 − 3)

Equation (2) can be re-written using 𝑘𝑘𝑑𝑑 and [𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑜𝑜𝑑𝑑,∞:

- 38 -

Equation (3-4) was applied to estimate the unknown constants, 𝑘𝑘𝑑𝑑 , 𝑘𝑘𝑟𝑟 𝑎𝑎𝑛𝑛𝑑𝑑 [𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑜𝑜𝑑𝑑,∞ from fits of

the dissolution data collected for the soils over time. Note that these three parameters are correlated

by Equation (3-3). The Euler method was applied to solve equation 3-3 numerically. 𝐾𝐾𝑓𝑓𝑓𝑓𝑙𝑙𝑓𝑓𝑓𝑓𝑑𝑑 was

estimated from the experimental data (Equation 3-5). From control experiments extracting Cu from

CuSO4 dosed soil, the efficiency of DTPA extraction, 𝜂𝜂𝐷𝐷𝑇𝑇𝐷𝐷𝐷𝐷 , was estimated to be 80%.

𝐾𝐾𝑓𝑓𝑓𝑓𝑙𝑙𝑓𝑓𝑓𝑓𝑑𝑑 =[𝐶𝐶𝐶𝐶]𝐷𝐷𝑇𝑇𝑃𝑃𝐴𝐴 𝜂𝜂𝐷𝐷𝑇𝑇𝑃𝑃𝐴𝐴

[𝐶𝐶𝐶𝐶]𝐶𝐶𝑙𝑙𝐶𝐶𝑙𝑙2∙𝑥𝑥𝐶𝐶𝐶𝐶2+ (3-5)

Where as [𝐶𝐶𝐶𝐶]𝐷𝐷𝑇𝑇𝐷𝐷𝐷𝐷 is DTPA extractable Cu, 𝜂𝜂𝐷𝐷𝑇𝑇𝐷𝐷𝐷𝐷 is the extraction efficiency (0.8 in this study),

[𝐶𝐶𝐶𝐶]𝐶𝐶𝑓𝑓𝐶𝐶𝑓𝑓2 is CaCl2 extractable Cu, and 𝑥𝑥𝐶𝐶𝐶𝐶2+ is the fraction of free Cu ions in soil pore water.

3.3 Results and Discussion

3.3.1 Effect of Soil Organic Matter on dissolution of CuO NP in soil.

To investigate the effect of SOM on dissolution of CuO NP in soil, a dissolution test in Lufa 2.1 soil

(100 mg/kg dw CuO NP treatment) and in Lufa 2.1 with added SOM (300 mg/kg dw CuO NP

treatment) was conducted (Figure 3-2). Different concentrations of CuO NP were applied based on

the estimated solubility from preliminary experiments (described in Appendix 2). Using the

dissolution model described in the methods section, the modeled solubility should increase from 95

mg/kg ( 95% CI: 87-108 mg/kg) to 254 mg/kg (95% CI: 234-280 mg/kg) in the amended soil

(Table 1). Doubling the reactive organic carbon content in Lufa 2.1 soil increased the solubility of

CuO NP by 2.7-fold, suggesting reactive organic carbon holds the main Cu pool in soil. Although

the solubility increased by 2.7-fold, the modeled dissolution rate constants between Lufa 2.1 soil and

Lufa 2.1 soil with added SOM are similar (95% confidence intervals are overlapping), suggesting that

SOM mainly affects the solubility of CuO NP in soil, but not its dissolution rate.

- 39 -

0 10 20 30 400

50

100

150

200

250

Time (days)

DTPA

Ext

ract

able

Cu

(mg

/kg

drie

d so

il) Lufa 2.1 soil, SOM added

Lufa 2.1 soil

Figure 3-2. Dissolution kinetics of CuO NP in Lufa 2.1 soil without added SOM (100 mg/kg dw

CuO NP treatment, circles) or with added SOM (300mg/kg dw CuO NP treatment, triangles). Bars

are standard deviation of the extractable Cu measurements (3 replicates). Soil pH in these studies

was 5.0 (unamended Lufa 2.1 soil) and 4.9 (Lufa 2.1. amended with SOM).

3.3.2 Effect of soil pH on dissolution of CuO NP in soil. The effect of soil pH on the dissolution behavior of CuO NP was investigated by modifying the pH

of Lufa 2.1 soil (100 mg/kg dw CuO NP treatment) and Lufa 2.2 soil (500 mg/kg dw CuO NP

treatment) with either CaO or CaCO3. Figure 3 indicates that higher pH significantly slowed down

the dissolution rate of CuO NP in soil in Lufa 2.2 soil. The modeled dissolution rate constant

decreased from 0.56 (CI95: 0.35-0.84)) (mg1/3·kg1/3·s-1) in Lufa 2.2 soil (pH=5.9) to 0.17 (CI95: 0.14-

0.21) (mg1/3·kg1/3·s-1) in Lufa 2.2 soil with pH adjustment (pH=6.8). For Lufa 2.1 soil (Appendix 2,

Figure A2-2), the dissolution of CuO NPs in pH-adjusted soil (pH=7.4) could not be accurately

modeled because of very limited dissolution, but it was clear that it was much slower than the

dissolution in Lufa 2.1 soil without pH adjustment (pH=5.0, kd= 0.83 mg1/3·kg1/3·s-1, with 95% CI:

0.65-1.00) during the 31d aging period. Although the dissolution rate constants are different,

suggesting a different particle lifetime in soil, the modeled solubility of CuO NPs in Lufa 2.2 soil

with and without pH adjustment are similar (Table 3-1). This can be observed from the extended

trend lines (dash lines) from the modeled dissolution kinetics in Figure 3-3. Thus, the soil pH mainly

determines how fast CuO NPs dissolve but has no measurable impact on their solubility. This is because most of

the Cu ions released from CuO NPs are retained by SOM. Carboxylic acid functional groups (pKa

- 40 -

<5) and weak acid groups (phenolics, pKa>9) mainly contribute to the acidity of humic acid (the

main component of SOM)50,51. The binding capacity between Cu and SOM is not sensitive to pH at

agriculture soil relevant pH (5 ~ 7.5)52 because the protonation state of SOM is not susceptible to

pH variation in this range. Thus, for a typical agriculture soil, although an increase in soil pH should

slow down the ion release process from CuO NP, it may have limited impact on the solubility of

CuO NP in that soil.

Figure 3-3. DTPA extractable Cu in Lufa 2.2 soil dosed with 500 mg/kg CuO NP at pH 5.9

(squares) and pH 6.8 (triangles). Dashed lines are model results showing the longer time trend. ‘X’ at

t=300 days is modeled maximum DTPA extractable Cu for each treatment. Bars are standard

deviation of the measurements (3 replicates) or the 95% confident intervals of the modeled

maximum DTPA extractable Cu (t= 300 day).

3.3.3 Effect of soil moisture content on the dissolution rate and solubility of CuO NP in soil.

As suggested from Figure 3-4, moisture content had no impact on the dissolution kinetics of CuO

NP. The modeled dissolution rate constants (kd and kr) and solubility [𝑪𝑪𝑪𝑪𝟐𝟐+]𝑻𝑻𝑻𝑻𝑻𝑻,∞ are the same for

CuO NP dissolving in soil with 10% moisture content or with 21% moisture content (Table 3-1).

This is consistent with the dissolution model that we proposed in which the soil pore water reaches

an equilibrium state with the soil solid matrix, where most dissolved Cu is retained by the soil solid

surfaces, not the soil pore water6,46. Thus, soil moisture should not affect the dissolution rate or

solubility of CuO NPs. It is acknowledged that we did not test extremes of dryness (e.g. moisture

content <<10%) because they do not represent normal agricultural soil conditions. CuO NP

dissolution could potentially be affected by extreme dryness due to the lack of water needed to

dissolve the CuO NPs. It should be noted that we also did not consider the flooded condition,

- 41 -

which would be relevant for crops like rice. In a flooded condition, where microbial processes can

deplete the soil of oxygen and lead to a strongly reducing condition, redox reactions could play a

role, affecting the dissolution of CuO NP.

0 10 20 30 400

100

200

300

400

DTPA extraction, Lufa 2.2 soil

Time (days)

Extra

ctab

le C

u(m

g /k

g dr

ied

soil) 21% moisture content

10% moisture content

a)

0 10 20 30 400

1

2

3

Time (days)

Extra

ctab

le C

u(m

g /k

g dr

ied

soil) 21% moisture content

10% moisture content

b) CaCl2 extraction, Lufa 2.2 Soil

Figure 3-4. Effect of moisture content on the dissolution kinetics of CuO NP in soil. (a) DTPA

Extractable Cu in Lufa 2.2, (b) CaCl2 extractable Cu in Lufa 2.2 soil at field capacity or 10%

moisture content. Bars are standard deviations of the measurements (3 replicates).

3.3.4 Dissolution rate and solubility of CuO NPs in soils with various properties.

To further investigate the effects of soil pH and SOM content, which are the more important

factors controlling the dissolution of CuO NP in soils, CuO NP dissolution tests in soils with

various soil properties, including Lufa 2.1, 2.2, 2.3 and 2.4 soils and pH/organic carbon amended

soils, were measured. The speciation of Cu in selected CuO NP dosed soils was confirmed by

XANES (appendix). The experimental conditions (soil properties, NP concentration) and fitted

dissolution model parameters (kd, kr and solubility) are shown in Table 3-1.

- 42 -

Table 3-1. Dissolution rate and solubility of CuO NP in a range of soils with various properties

Soil Moisture content

Concentration of CuO NP (mg/kg

dw)

Organic carbon content (% C)

pH kd (mg1/3·kg1

/3·s-1)

kr (mg-

2/3·kg-2/3·s-

1)

Solubility (mg/kg)

R2

Lufa 2.1 soil

16% 115±7 0.67 5.0 0.83 (0.65-1.00)

0.19 (0.13-0.26)

95 (87-108)

0.992

Lufa 2.2 soil

21% 503±17 1.71 5.9 0.70 (0.55-0.92)

0.23 (0.15-0.33)

328 (300-370)

0.993

Lufa 2.4 soil

22% 537±89 1.99 7.2 0.12 (0.08-0.17)

0.09 (0.00-0.22)

315 (170-*1)

0.991

Lufa 2.3 Soil

17% 539±3 0.66 6.5 0.15 (0.11-0.21)

0.26 (0.15-0.38)

84 (75-99)

0.987

Lufa 2.2 soil- pH adjusted

21% 501±11 1.71 6.8 0.17 (0.14-0.21)

0.05 (0.02-0.08)

295 (220-550)

0.998

Lufa 2.2 soil

21% 265±13 1.71 5.8 0.56 (0.35-0.84)

0.19 (0.01-0.41)

319 (220-3800)

0.986

Lufa 2.2 soil

10% 481±24 1.71 5.9 0.57 (0.43-0.72)

0.20 (0.13-0.29)

325 (287-405)

0.994

Lufa 2.1 soil- pH

adjusted2

16% 112±5 0.67 7.4 N.A. N.A. N.A. N.A.

Lufa 2.1 soil- OM adjusted

16% 287±16 1.34 4.9 0.99 (0.84-1.18)

0.05 (0.03-0.06)

254 (234-280)

0.996

1: The upper bond of solubility cannot be determined with confidence.

2:Dissolution was too low to be modeled with confidence.

- 43 -

To investigate the correlations between soil properties (soil pH, organic carbon content) and

dissolution kinetics (solubility and dissolution rate constant), we conducted a multivariate regression

(Appendix 2 table A2-5 to-A2-7). The result showed that the solubility was correlated with soil organic

carbon content, but not {H+} whereas dissolution rate constant was correlated with {H+} but not organic carbon

content. A strong correlation between organic carbon content and solubility of CuO NP in soil

(Figure 3-5a, R2 = 0.89, 𝑁𝑁 < 0.0001, Solubility = 1.83 ∗ 102 ∗ 𝑂𝑂𝐶𝐶%) was observed, suggesting

soil organic carbon content is the main driver for the solubility of the CuO NPs in soil across soil

types.. It should be noted that our study did not cover soils with extreme high organic carbon

content (~10%). The SOM in our study ranged from ~0.8% to 3.5% (0.5% to 2% organic carbon

content). Thus, the developed correlation covers the low end to an intermediate range for

agricultural soils, but is likely to be predictive of 10% SOM (2.9x higher than our high end). On the

other hand, hydrogen ion activity, {H+}, is positively correlated with the dissolution rate constant

(Figure 3-5b). According to previous studies in aqueous systems, pH should play an important role

in the dissolution rate of CuO NP13,53,54. However, because most previous studies used empirical first

order dissolution models, the reaction mechanism and reaction order of CuO NP dissolution with

respect to hydrogen concentration remained unknown. In this study, no assumption about the

reaction order with respect to {H+} was made. Instead, reaction order was experimentally

determined by plotting the dissolution rate constant against {H+}. Figure 3-5 b shows the linear

relationship between the dissolution rate constant and {H+} for soil pH below 6.3 (R2 = 0.89,𝑁𝑁 =

0.016 𝑘𝑘𝑑𝑑 = 2.83 ∗ 104 ∗ {𝐻𝐻+} + 0.57 ). However, above pH=6.3 this relationship no longer

holds, and the dissolution rate constant is approximated by a single value (𝑘𝑘𝑑𝑑,𝑓𝑓𝑎𝑎𝑎𝑎𝑟𝑟𝑓𝑓𝑙𝑙𝑎𝑎 =

0.14, 95% 𝐶𝐶𝐼𝐼(0.07 − 0.21). Due to the limited amount of dissolution rate and solubility data, a

cross-validation was done, suggesting the fit was stable among the different soils (i.e., no particular

soil treatment inordinately affected the correlation, see Figure A2-3, Appendix 2). The pH-

dependent correlations suggest that the dissolution of CuO NP in soil with pH below 6.3 may be

governed by a different dissolution mechanism (Equation 3-6) compared to CuO NP dissolution in

soil with pH above 6.3 (Equation 3-7). At higher pH, the activity of H+ decreases, thus Equation 3-7

could be the dominant reaction pathway rather than Equation 3-6. Regarding the reverse reaction

(precipitation) process, no trend was found between kr and {H+} and organic carbon content (Table

A2-7). However, kr was correlated with kd and solubility by Equation (3-2) and is thus affected by

both the pH and SOM simultaneously.

- 44 -

CuO (s)+2H+(aq) ↔ Cu2+

(aq) +H2O(l) (3-6)

CuO(s) + H2O(l) ↔ Cu2+(aq)+ OH- (3-7)

0 . 0 0 . 5 1 . 0 1 . 5 2 . 0 2 . 5

0

2 0 0

4 0 0

6 0 0

O r g a n i c c a r b o n c o n t e n t ( % C )

So

lub

ilit

y (

mg

/kg

)

a )* *

Figure 3-5. Correlation between organic carbon content and solubility (a) and between {H+}and

dissolution rate constant, kd (b). The right figure in b) shows the high soil pH data in the red box. *:

upper 95% CI was high, see Table 3-1.

b)

- 45 -

3.3.5 Predicting CuO NP solubility and dissolution rate in two test soils. Using the correlations derived from Figure 3-5 a and b, we estimated the solubility and dissolution

rate constant of the CuO NPs in soil samples from Arizona and from Portugal (soil properties can

be found in appendix, table A2-1) based on their pH and soil organic carbon content. Note that the

95% CI of the prediction is calculated using uncertainties in estimating kd and solubility from the

correlations shown in Figure 3.5-a) and b). According to the correlations that we developed, the

predicted solubility of CuO NP in the Arizona soil is 96 ±13 mg/kg, with the first order dissolution

rate constant, 0.14±0.07 mg1/3·kg1/3·s-1, whereas the predicted solubility and first order dissolution

rate constant of CuO NP in the Portugal soil are 213±30 mg/kg and 0.68±0.12 mg1/3·kg1/3·s-1

respectively. The pH of the Arizona soil was in the high pH region (where kd is constant), whereas

the pH of the Portugal soil, being in the lower pH region where there is a linear relationship

between kd and pH. To determine the precision of the model predictions, a Monte Carlo simulation

(500 simulations based on the standard deviation and mean of the estimated kd and solubility to

generate a confidence interval) was used to simulate the dissolution kinetics of CuO NP in Arizona

soil and in Portugal soil (Figure 3-6). In both cases, the experimental data fell within 95% CI. This is

evidence that the correlations that we developed can predict the dissolution rate and solubility of

CuO NP in Arizona soil and in Portugal soil based on its pH and organic carbon content. The

experimentally measured CuO NP dissolution was lower than the best fit model prediction in the

Arizona soil and was higher than the best fit model prediction in the Portugal soil. This discrepancy

may be due to our assumption that the composition of SOM and DOM was the same among

different soils. These assumptions contribute to the uncertainties in predicting the solubility and the

dissolution rate constant of CuO NPs in soil. This is because the different chemical composition of

SOM and DOM in various soils may indeed affect the ability of SOM/DOM to complex with Cu or

affect its dissolution rate constant.13

0 1 0 2 0 3 0 4 0

0

5 0

1 0 0

1 5 0

T i m e ( d a y s )

Dis

so

lve

d C

u (

mg

/kg

)

( a )

0 1 0 2 0 3 0 4 0

0

1 0 0

2 0 0

3 0 0

4 0 0

T i m e ( d a y s )

Dis

so

lve

d C

u (

mg

/kg

)

E x p e r i m e n t a l d a t a

P r e d i t i o n h i g h e r 9 5 % C I

P r e d i t i o n l o w e r 9 5 % C I

P r e d i c t e d d i s s o l u t i o n k i n e t i c s

( b )

- 46 -

Figure 3-6. Prediction (predicted dissolution kinetics is the red line, upper 95% CI is the blue line,

lower 95% CI is the purple line) and experimental data (triangle, Cu% dissolved estimated from

DTPA extraction) of CuO NP dissolution in an Arizona soil (a) and in a Portugal soil (b).

3.4 Environmental Implications Previous studies proposed that soil pH is an important factor affecting the dissolution rate of CuO

NP7,8. In this paper, we mechanistically demonstrate that, while soil pH mainly affects how fast the

CuO NPs dissolve, SOM content is the main factor affecting the solubility of CuO NPs in soil (how

much of the NPs can dissolve at equilibrium). The dissolution rate and solubility of CuO NPs

together describe their overall dissolution kinetics in soil. We also found that the soil moisture

content had no impact on the dissolution kinetics of CuO NPs in soil due to the low mass of Cu in

porewater compared to the other soils sinks for Cu (e.g. SOM and mineral surfaces).

With the dissolution model and correlations developed in this study, the dissolution kinetics of the

studied CuO NPs can be predicted from the SOM content and soil pH. This enables modelers to

include the dissolution process of the CuO NP (with certain size, shape and make) in their model

without the need for experimental data or guessing55. Our study also predicted the dissolution rate

and solubility of CuO NPs, which is needed to evaluate their environmental risks56. However, these

results are specific to the CuO NPs studied here. Work is still needed to investigate how the size,

shape and coatings affect the dissolution kinetics of CuO NP in soil. It is possible to incorporate

these properties in the dissolution model. For example, we represent the shape of the particle by the

shape factor, n (Appendix 2, eq A2-3). For a spherical particle or a cube, n=2/3, but for particles

with other shapes, the shape factor would change. The initial size of the particle would affect the

conversion from particle mass concentration to total surface area, which is reflected by C1

(Appendix 2, eq A2-3). The model that we developed in this study for CuO NPs can be potentially

extended to model the dissolution rates of other metal-based ENMs when the time scale for

dissolution is much longer than the time scale of the sorption behavior of ions to soil surfaces. It

should be noted that pH and the amount of SOM were found to be the most important soil

properties affecting the dissolution of CuO NP in soil, but other soil properties could be important

as well. A better understanding of how the composition of SOM and DOM affects the dissolution

of CuO NP could make the model more accurate. With more specific characterization of SOM and

DOM composition, it may indeed be possible to improve the predictive capability of the models.

However, it would require more detailed characterization of SOM and DOM composition as a

- 47 -

trade-off. Our study did not eliminate microbial activity in soil, which could also contribute to the

dissolution of ENMs in soil57,58. However, on a gross level, the impacts of microbial activity seem to

be small relative to the impacts of pH and SOM content. Future models could distinguish between

the microbial and geochemical contributions to dissolution. Our previous study also suggested that

root activity could play a role in dissolution of CuO NP in rhizosphere soil10. Our current model did

not incorporate such influences, but this could be an interesting topic for future research.

Previous studies showed that SOM is the biggest Cu sink in soil, but other soil surfaces, like clay and

iron hydroxide surfaces could be more important sinks for other metals47,49. Our study

demonstrated that, in addition to nanoparticle properties, soil properties should be considered when

predicting the risks or efficiency of ENMs being applied to soil.

This study showed that, unlike the dissolution kinetics in aqueous systems, in soil, pH affects the

dissolution rate constant for CuO NPs, but not their overall solubility . Instead, soil organic matter,

which provides the sink for the dissolved Cu species, was found to control the overall solubility of

CuO NP in soil.

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(11) McManus, P.; Hortin, J.; Anderson, A. J.; Jacobson, A. R.; Britt, D. W.; Stewart, J.; McLean, J. E. Rhizosphere interactions between copper oxide nanoparticles and wheat root exudates in a sand matrix: Influences on copper bioavailability and uptake. Environ. Toxicol. Chem. 2018, 37 (10), 2619–2632.

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(18) Sekine, R.; Brunetti, G.; Donner, E.; Khaksar, M.; Vasilev, K.; Jamting, A. K.; Scheckel, K. G.; Kappen, P.; Zhang, H.; Lombi, E. Speciation and Lability of Ag-, AgCl-, and AgS-Nanoparticles in Soil Determined by X-ray Absorption Spectroscopy and Diffusive Gradients in Thin Films. Env. Sci Technol 2014.

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(20) Pradas del Real, A. E.; Castillo-Michel, H. A.; Kaegi, R.; Sinnet, B.; Magnin, V.; Findling, N.; Villanova, J.; Carriere, M.; Santaella, C.; Fernandez-Martinez, A. Fate of Ag-NPs in sewage

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sludge after application on agricultural soils. Env. Sci Technol 2016.

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(23) Peijnenburg, W.; Jager, T. Monitoring approaches to assess bioaccessibility and bioavailability of metals: matrix issues. Ecotoxicol. Environ. Saf. 2003, 56 (1), 63–77.





(28) Degryse, F.; Broos, K.; Smolders, E.; Merckx, R. Soil solution concentration of Cd and Zn canbe predicted with a CaCl2 soil extract. Eur. J. Soil Sci. 2003, 54 (1), 149–158.


(30) Kashem, M. A.; Singh, B. R.; Kondo, T.; Huq, S. M. I.; Kawai, S. Comparison of extractability of Cd, Cu, Pb and Zn with sequential extraction in contaminated and non-contaminated soils. Int. J. Environ. Sci. Technol. 2007, 4 (2), 169–176.

(31) Sahuquillo, A.; Rigol, A.; Rauret, G. Overview of the use of leaching/extraction tests for risk assessment of trace metals in contaminated soils and sediments. TrAC Trends Anal. Chem. 2003, 22 (3), 152–159.

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(33) Husson, O. Redox potential (Eh) and pH as drivers of soil/plant/microorganism systems: a transdisciplinary overview pointing to integrative opportunities for agronomy. Plant Soil 2013, 362 (1–2), 389–417.

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34 (7), 1125–1131.

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(37) Karlsson, T.; Persson, P.; Skyllberg, U. Complexation of copper (II) in organic soils and in dissolved organic matter− EXAFS evidence for chelate ring structures. Environ. Sci. Technol. 2006, 40 (8), 2623–2628.

(38) Sheoran, V.; Sheoran, A. S.; Poonia, P. Soil reclamation of abandoned mine land by revegetation: a review. Int. J. Soil, Sediment Water 2010, 3 (2), 13.

(39) Nelson, D. W.; Sommers, L. E. Total carbon, organic carbon, and organic matter. Methods soil Anal. part 3—chemical methods 1996, No. methodsofsoilan3, 961–1010.

(40) van Zomeren, A.; Comans, R. N. J. Measurement of humic and fulvic acid concentrations and dissolution properties by a rapid batch procedure. Environ. Sci. Technol. 2007, 41 (19), 6755–6761.

(41) Dijkstra, J. J.; Meeussen, J. C. L.; Comans, R. N. J. Evaluation of a generic multisurface sorption model for inorganic soil contaminants. Environ. Sci. Technol. 2009, 43 (16), 6196–6201.

(42) Gao, X.; Spielman-Sun, E.; Rodrigues, S. M.; Casman, E. A.; Lowry, G. V. Time and Nanoparticle Concentration Affect the Extractability of Cu from CuO NP-Amended Soil. Environ. Sci. Technol. 2017, 51 (4).

(43) Lierop, W. van. Conversion of organic soil pH values measured in water, 0.01 M CaCl2 or 1 N KCl. Can. J. Soil Sci. 1981, 61 (4), 577–579.

(44) Eckert, D.; Sims, J. T. Recommended soil pH and lime requirement tests. Recomm. soil Test. Proced. Northeast. United States. Northeast Reg. Bull. 1995, 493, 11–16.

(45) Degryse, F.; Smolders, E.; Parker, D. R. Partitioning of metals (Cd, Co, Cu, Ni, Pb, Zn) in soils: concepts, methodologies, prediction and applications - a review. Eur. J. Soil Sci. 2009, 60 (4), 590–612.



(48) Donner, E.; Ryan, C. G.; Howard, D. L.; Zarcinas, B.; Scheckel, K. G.; McGrath, S. P.; de Jonge, M. D.; Paterson, D.; Naidu, R.; Lombi, E. A multi-technique investigation of copper and zinc distribution, speciation and potential bioavailability in biosolids. Env. Pollut 2012, 166, 57–64.

(49) Duffner, A.; Weng, L.; Hoffland, E.; van der Zee, S. E. Multi-surface modeling to predict free zinc ion concentrations in low-zinc soils. Env. Sci Technol 2014, 48 (10), 5700–5708.

(50) Jeong, C. Y.; Park, C. W.; Kim, J.-G.; Lim, S. K. Carboxylic content of humic acid determined

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by modeling, calcium acetate, and precipitation methods. Soil Sci. Soc. Am. J. 2007, 71 (1), 86–94.

(51) Ritchie, J. D.; Perdue, E. M. Proton-binding study of standard and reference fulvic acids, humic acids, and natural organic matter. Geochim. Cosmochim. Acta 2003, 67 (1), 85–96.

(52) Villaverde, P.; Gondar, D.; Antelo, J.; Lopez, R.; Fiol, S.; Arce, F. Influence of pH on copper, lead and cadmium binding by an ombrotrophic peat. Eur. J. Soil Sci. 2009, 60 (3), 377–385.

(53) 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 (10), 375–380.


(55) Dale, A. L.; Casman, E. A.; Lowry, G. V; Lead, J. R.; Viparelli, E.; Baalousha, M. Modeling nanomaterial environmental fate in aquatic systems. Env. Sci Technol 2015, 49 (5), 2587–2593.

(56) Utembe, W.; Potgieter, K.; Stefaniak, A. B.; Gulumian, M. Dissolution and biodurability: Important parameters needed for risk assessment of nanomaterials. Part. Fibre Toxicol. 2015, 12 (1), 11.

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CHAPTER 4: Elucidate CuO NP dissolution behavior and toxicity to wheat (Triticum aestivum) in rhizosphere and bulk soil.

Abstract: It has been suggested, but not previously measured, that dissolution kinetics of soluble nanoparticles

such as CuO NPs in soil affect their phytotoxicity. An added complexity is that such dissolution is

also affected by the presence of plant roots. Here, we measured the rate of dissolution of CuO NPs

in bulk soil, and in soil in which wheat plants (Triticum aestivum) were grown under two soil NP

dosing conditions: (a) freshly added CuO NPs (500 mg Cu/kg soil), and (b) CuO NPs aged for 28d

before planting. At the end of the plant growth period (14 days), available Cu was measured in three

different soil compartments: bulk (not associated with roots), loosely attached to roots, and

rhizosphere (soil firmly attached to roots). The labile Cu fraction increased from 17mg/kg to

223mg/kg in fresh treatments and from 283 mg/kg to 305mg/kg in aged treatments over the

growth period due to dissolution. Aging CuO NPs increased the toxicity to Triticum aestivum

(reduction in root maximal length). The presence of roots in the soil had opposite and somewhat

compensatory effects on NP dissolution, as measured in rhizosphere soil. pH increased 0.4 pH units

for fresh NP treatments and 0.6 pH units for aged NPs. This lowered CuO NP dissolution in

rhizosphere soil. Exudates from T. aestivum roots also increased soluble Cu in porewater. CaCl2

extractable Cu concentrations in bulk vs. rhizosphere soil increased from 1.8mg/kg to 6.2mg/kg

(fresh treatment), and from 3.4mg/kg to 5.4mg/kg (aged treatments). Our study correlated CuO NP

dissolution and the resulting Cu ion exposure profile to phytotoxicity, and showed that plant-

induced changes in rhizosphere conditions should be considered when measuring the dissolution of

CuO NP near roots.

This work has been published in Environmental Science & Technology as ‘CuO nanoparticle

dissolution and toxicity to wheat (Triticum aestivum) in rhizosphere soil’ , doi: acs.est.7b05816

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4.1 Introduction

The anticipated benefits of nano-enabled agrochemicals include slow and controlled release of

micronutrients, plant tissue-specific targeted release of micronutrients or pesticides, reduced

amounts of agrochemicals being required, and generally lower toxicity compared to more soluble

products1,2. Copper-based nanoparticles (NPs) are already on the agrochemical market3,4. Copper is

an essential crop micronutrient. Deficiency may lead to reduced disease resistance5 and decreased

crop yields6. However, at high concentrations, Cu can also be toxic to plants,7 the surrounding

microbial communities,8 and soil invertebrates9. Due to its relatively slow dissolution, CuO NPs

have been studied as a potential candidate for agrochemical use. It behaves differently from

dissolved Cu2+ in soil, potentially affecting copper bioavailability, the release of Cu ions over time,

and potential associated risks10–12. However, the connection between NP dissolution, the resulting

dose of Cu ions and its toxicity to terrestrial plants, and the role of root exudates on this process

have not been well elucidated due to a lack of appropriate characterization of the dissolution of the

NPs in soil. Ideally, application rates of these novel materials should be based on their fate and

effects in the terrestrial environment, their bioavailability and potential toxicity to plants. The toxic

effect of Cu species is reflected in physiological changes in plant roots and shoots, such as decreased

root length, increased root compactness, change in root color, shorter leaf length and decreased

shoot biomass13–15. Hyperspectral imaging has been used to visualize NPs in plants and to confirm

macroscopic evidence of NP toxicity16,17.

Previous studies of the toxicity of CuO NPs to terrestrial plants assumed, but did not

measure, dissolution behavior of CuO NP in soil. This has led to conflicting conclusions on the

toxicity of CuO NPs. While some studies attributed the toxic effect of CuO NP to released ionic

Cu15,18,19, others concluded the opposite20. For example, Servin et al. chose a Cu ion control

concentration based on the assumption that only 10% of the CuO NP would dissolve in soil, the

same fraction that dissolved in pure sand, rather than measuring CuO dissolution in soil. They

concluded that dissolution of CuO NPs could not fully explain the plant toxicity because the plant

responses differed from their Cu ion control .20 Much more than 10% CuO NP could have

dissolved in soil because soil organic matter (SOM) acts as a Cu sink, increasing the amount of CuO

NP that can be dissolved11. This weakens their conclusions about a NP-specific effect. Similar

problems occurred in other studies 15,18,21–23. Breaking with this trend, Dimkpa et al. (2013) evaluated

the total CuO NP dissolved in soil using a water-extraction method. 24 Unfortunately, the water-

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extraction does not extract Cu bound to the soil solid matrix which accounts for most of the

dissolved Cu in soils25–27. Thus, their assertion of a CuO NP-specific toxicity in soil is confounded by

the potential that more Cu had dissolved than was assumed or measured. Recently Qiu et al. found

that the toxicity of CuO NP, CuO bulk particles and soluble Cu (Cu(AC)2) depends on their

solubility in soil, and that the distinction in solubility diminished after a 90-day aging period.

However, the actual dissolution during the incubation periods (one day vs. 90 day) was not

quantified. They successfully correlated the toxicity of NP to roots of Hordeum vulgare L. (5-day root

elongation experiment) with ‘free Cu ions’ in soil pore water measured at a single time point before

seeding;19 though convincing, it should be noted that the dissolution during the 5-day toxicity test

was not considered. While the relatively slow dissolution of CuO NP may result in unobservable

impacts on toxicity during a relatively short 5-day toxicity test, dissolution at this rate would

probably affect toxicity of NPs in longer tests.

The dissolution of CuO NPs is a dissolution rate-limited process. Experimental approaches,

such as extraction with CaCl2 or with diethylenetriaminepentaacetic acid (DTPA), have been used to

predict the bioavailability or toxicity of metals in soil.28–31 CaCl2 extracts the Cu ions in soil pore

water that are considered ‘readily available,’ while DTPA extracts the “labile” fraction including

dissolved Cu in soil pore water (free Cu2+and Cu2+ complexed with soluble ligands such as dissolved

organic matter (DOM)), but also the Cu2+ associated with soil solid phases, such as soil organic

matter (SOM), clay particles, and iron oxide minerals.29–31 Whereas CaCl2 extracts metals that are

‘readily available’ to plants29, DTPA extracts this pool as well as the pool that may eventually

become bioavailable in soil, the so called ‘potentially available’ fraction32. One problem with using

these extraction methods to predict the bioavailability of Cu based nanomaterials is that a single time

point extraction does not capture the temporal dynamics of the CuO NP dissolution process. Our

recent study used extraction methods at different times to monitor the kinetics of release of Cu ions

from CuO NP in soil. In that study, the increase in DTPA extractable Cu over 30 days in soils was

used to estimate the dissolved pool of Cu in soil.11 The availability of Cu ions increased with time

over a 30d period, which may explain why previous efforts to correlate the extractable metals in

metal-based NP-amended soils with their bioavailability or toxicity have generally failed33–35.

Plants also may affect the dissolution behavior and availability of CuO NP in soil, especially

in the rhizosphere. Previous studies using extraction methods to predict the bioavailability or toxicity

of metal-based ENMs or the dissolution of ENMs in soil did not typically consider the impact of

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roots on Cu availability.11,12,33–35 Plant roots exude organic acids 36–38 that may affect the pH in

rhizosphere. 39,40 Although soil pH and organic carbon are known to be important factors

influencing the dissolution behavior of CuO NPs in soil11,12, and previous studies have proposed that

exudates from plant roots may affect the dissolution of CuO NP in the rhizosphere41, no studies

have quantified this. Given that the rhizosphere is where plants interact with soil for nutrient

uptake,42,43 a better understanding of how the roots impact NP dissolution and metal availability in

the rhizosphere is needed to design nano-enabled agrichemicals with optimal properties for

delivering nutrients.

The objectives of this study are to quantify the influence of time and near-root chemical

conditions on dissolution and lability of CuO NPs in rhizosphere soil, and to determine the

influence of this dissolution on the toxicity of CuO NPs to Triticum aestivum during a 14-day plant

growth period in soil. Wheat (Triticum aestivum) was used in this study because it is the 2nd most

cultivated plant in the world, and it is sensitive to Cu deficiency44 or excess45. To evaluate the toxicity

of CuO NP to plants, we measured the dissolution behavior of CuO NPs in soil in the presence of

plants with emphasis on the soil-plant interface (rhizosphere) where roots interact with soil. The

toxicity of Cu was evaluated by physiological changes in plant roots and shoots.


4.2.1 Chemicals

Calcium chloride (≥99.0%, ACS grade), calcium oxide (CaO), calcium carbonate (CaCO3) (99%+),

and hydrogen peroxide (30%, certified ACS) were purchased from Fisher Scientific. DTPA (>99%)

and triethanolamine (TEA, ≥99.0% (GC)) were purchased from Sigma-Aldrich. Trace metal grade

nitric acid (65%-70%) was purchased from VWR. Copper sulfate (CuSO4) was purchased from

Fisher Scientific . Lufa Standard soils (2.1, 2.2, 2.4 and 2.4) were purchased from Lufa Speyer,

Germany. A calcareous soil (pH 7.6) was collected in Arizona (termed Arizona soil) and used to test

the model’s ability to predict CuO NP dissolution behavior based on soil pH and SOM content.

Another more acidic soil (pH=5.0) was collected from a grassland in northwestern Portugal (termed

Portugal soil). Detailed properties of all the soils used can be found in appendix ( Table A3-1).

Calcium chloride (≥99.0%, ACS grade) and hydrogen peroxide (30%, certified ACS) were

purchased from Fisher Scientific. DTPA (>99%) and triethanolamine (TEA, ≥99.0% (GC)) were

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purchased from Sigma-Aldrich. Trace metal grade nitric acid (65%-70%) was purchased from VWR.

Triticum aestivum seeds (Pembroke 2014) were bred by Dr. David Van Sanford (Department of Plant

and Soil Sciences, University of Kentucky).

4.2.2Nanoparticle Characterization

CuO NPs (~40 nm primary particle size), were purchased from Sigma-Aldrich. The primary size of

particles, zeta potential, isoelectric point and hydrodynamic diameter have been characterized and

reported in Chapter 26.

4.2.3 Soils and Characterization of Soil Properties Standard Lufa 2.2 soil (loamy sand) was purchased from Lufa Speyer, Germany. Lufa 2.2 soil

contains 1.6 wt. % soil organic matter, and little total and available Cu (see Appendix 3, Table A3-1

and Table A3-2, control treatment). Using a well-characterized standard soil allows comparisons

between studies. The high carbon organic content (about 1.6%) of Lufa 2.2 makes this soil good for

agricultural studies. Soil was air dried and sieved < 2mm before shipping. The soil was further air-

dried for at least 24 hours before all experiments. Soil pH in different treatments was determined by

the CaCl2 extraction method (see ‘Extraction methods’ section). Soil moisture content (1% for the

air dried soil) was determined gravimetrically after oven-drying the soil at 105 ºC for 24 h46. Soil field

moisture capacity (21%) was determined using a Haines apparatus with 0.1 bar pressure difference

between the wet soil and the atmosphere. Soil was in equilibrium with air and not water saturated,

and presumed to be aerobic for the duration of the experiment (~+400mV)

4.2.4 Soil amendment. The CuO NP suspension (containing Na2SO4), CuSO4 solution, or Na2SO4 solution (control

treatment) were mixed with soil and brought to a moisture content of 21.7% (corresponding to

~50% of the water holding capacity). The soil was mixed with wooden sticks in a beaker for 20min.

The homogeneity was confirmed with the low standard deviation for the total Cu content measured

by soil digestion data (Appendix 3, table A3-1). To test if CuO NP and CuSO4 treatments resulted in

different Cu bioavailability and toxicity, the Cu ion concentration had to be high enough to ensure

some CuO NPs remained in the soil during the study period. We chose 500mg/kg (as Cu) for the

CuO NP treatment, and 300 mg/kg (as Cu) for the CuSO4 treatment based on a preliminary study

to assess the solubility of the CuO NPs in the Lufa 2.2 soil (Appendix 3, Figure A3-1). The results

showed that the solubility of CuO NPs in Lufa 2.2 soil was ~300mg/kg. Therefore, the selected

- 57 -

concentrations provided a similar concentration of added Cu ion in both treatments after one

month.

4.2.5 Germination and plant growth

The seeds of Triticum aestivum were surface sterilized by submerging them in 10% sodium

hypochlorite solution for 10 minutes, and then washed with DI water three times. The seeds were

then kept immersed in DI water overnight on an end-to-end rotator. The following day, the seeds

were transferred to a petri-dish containing moist tissue paper. The petri-dishes were covered with

aluminum foil and incubated in the growth chamber for 7 days, until 90% germination was achieved.

Germinated seeds were transplanted into syringes containing 120g of amended soils either

immediately after adding the Cu (fresh treatment) or 28 days after the Cu was added (aged

treatment).The plants were incubated in a growth chamber with constant moisture content and 16h-

light/8h-dark cycle (25 °C for daytime and 21 °C for night time). A diluted Cu-free Hoagland

solution (quarter strength) was added (1ml/day) to each syringe to maintain the moisture content of

the soil as well as provide nutrients to plants. The concentration of Cu in soil and plant tissue was

determined using a standard digestion method (EPA Method 3050b47) and ICP-MS analysis of Cu in

the digestate. See appendix. Adding moisture content did not induce any vertical transport of Cu, as

suggested by Figure A3-4 in Appendix 3.

4.2.6 Sampling of soil and plant tissue Prior to transplanting the germinated seeds in soils, subsamples of each soil were collected from all

treatments for DTPA extraction (2g of soil per extraction) to measure the labile metal fraction. After

14d of growth, rhizosphere soil, "loosely attached soil," and bulk soil (Appendix 3, Figure A3-2)

were collected for DTPA and CaCl2 extraction to determine the total dissolved metal and readily

available metal, respectively, as described below. After the plants and roots were removed from the

syringe, the soil remaining inside the syringe was defined as bulk soil, presumably minimally affected

by the plant roots. The bottom 5mm of bulk soil was also collected to determine if there was

significant vertical transport of Cu. No vertical transport of Cu was observed (Appendix 3, Figure

A3-4). The roots were separated from shoots. Both roots and shoots were photographed with a

scale bar for determination of length. For each treatment, one plant root replicate was washed with

1mM KCl three times for Cytoviva analysis (described below). The remaining roots were shaken by

hand in a 50 ml centrifuge tube, and the soil that detached during shaking was defined as loosely

attached soil48 (Appendix 3, Figure A3-2). After shaking, the roots were placed on aluminum foil and

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air dried in a fume hood for 24 hours. The roots were then shaken again in a 50 ml centrifuge tube,

and the soil that detached during the air-drying process and the second shaking process was defined

as rhizosphere soil49. Due to the small amount of rhizosphere soil collected per treatment, not all

replicates were suitable for DTPA and CaCl2 extraction. The details of which samples were analyzed

can be found in appendix (Table A3-3 and Table A3-4). For CuSO4 treatments, the roots were

highly compacted, precluding the collection of rhizosphere soil. The shoots and roots were oven-

dried at 105 ºC for 24 h. The mass of the dried roots and shoots was recorded before digestion for

total Cu analysis (details in Appendix C).

4.2.7 Soil extraction

Two standard extraction fluids were used in this study. DTPA extractant was composed of 0.01M

CaCl2, 0.005M DTPA and 0.1M triethanolamine (TEA) (pH=7.6). CaCl2 extractant was 0.01M CaCl2

without pH adjustment. All extractions were done using a reciprocal shaker at 180 rpm for 2 hours.

It is important to note that the centrifuge tubes were laid horizontally in the shaker rather than

vertically to provide the best extraction efficiency. For soil samples collected before the plant growth

experiments, 2g of soil were extracted with 4ml DTPA extractant. For bulk soil samples, loosely

attached soil samples and rhizosphere soil samples, 0.4g of soil was extracted with 0.8ml DTPA

extractant, while 0.35g of soil was extracted with 3.5ml CaCl2 extractant. After extraction, all samples

were centrifuged at 3000 rpm for 10 min, and the supernatants were filtered using a 0.45 µm PTFE

filter. The pH of the CaCl2 extracts for each soil fraction was measured (Figure 5). All samples were

acidified with 20% HNO3 (final HNO3 concentration, 2%) and Milli-Q-water before being analyzed

by ICP-MS. The method for ICP-MS is provided in detail in the appendix C.

4.2.8 Cytoviva analysis

The interaction between roots and NPs were visualized in fresh roots after a rinsing step in 10-3 M

KCl, using a darkfield-based hyperspectral imaging (DF-HSI) system (CytoViva Inc., USA). See

appendix C for additional details.

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4.3 Results and Discussion

4.3.1 Nanoparticle characterization

The properties of the CuO NPs have been previously described11. Briefly, the primary particle size

was determined by TEM to be 38nm ± 1.7nm (278 particles counted, 95% CI). The hydrodynamic

diameter of an 80mg/kg CuO NP in an aqueous 5mM NaHCO3 suspension (pH=7) was measured

to be 560nm±103nm (3 replicates, 95% CI, intensity averaged), and the zeta potential was -

16.1mV±1.7mV (3 replicates, 95%CI) in the same suspension. The pH of the isoelectric point

(pHiep) of the CuO NPs in a 5mM NaNO3 solution was 8.8, while the pHiep was 5.8 in the 5mM

NaHCO3 solution11. The hydrodynamic diameter and zeta potential likely change after they are

added to the soils due to interactions with soil components such as natural organic matter and

calcium50,51.

4.3.2 Change in extractability of Cu in bulk soil during the plant growth experiment.

The DTPA extractable Cu in the bulk soil for CuO NP and CuSO4 amended soils are shown in

Figure 4-1. The DTPA extractable Cu represents the Cu that was released from the CuO NPs

during the treatment. The extractable Cu vs. time is shown for the 14d growth period for both the

freshly added CuO NPs, and for the aged CuO NPs, where plants were added after the CuO NPs

had aged for 28d prior to planting the germinated seeds. The total Cu concentration in the two

treatments and in the control (unamended) soil is provided in the appendix (Table A3-1).

For the CuO NP treatments, the DTPA extractable Cu from bulk soil increased over time

(Figure 4-1a) (ANOVA test, P<0.05), although the increase was much higher for fresh CuO NP

treatment than that for the aged CuO NP treatment. The DTPA extractable Cu in bulk soil collected

form these plants was measured at four points (just after mixing, t=14, 28, and 42d) during the

study. The dissolution rate fit a first-order dissolution model well (R2=0.990). The aged CuSO4

treatments showed the opposite trend, with DTPA extractable Cu slightly decreasing over time

(Figure 4-1b). The DTPA extractable Cu in soils from the control treatment (Na2SO4) was low, and

no change was observed during the 2-week plant growth experiments (Appendix 3, Table A3-2).

At the end of the 14d plant growth period for the fresh CuO NP treatment, the DTPA extractable

Cu was similar to the aged CuSO4 treatments (ANOVA test, P>0.05). The DTPA extractable Cu in

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the aged CuO NP treatment was statistically significantly higher than both fresh and aged CuSO4

treatment (ANOVA test followed by Fisher’s LSD test for multiple comparison, P≤0.05) (Figure 4-

1c). The CaCl2 extractable Cu revealed a different order, with fresh CuSO4 treatment having the

highest CaCl2 extractable Cu, followed by the aged CuO NP treatment and the aged CuSO4

treatment, with the fresh CuO NP treatment having the lowest amount of CaCl2 extractable Cu. The

CaCl2 extractable copper represents the “readily available” Cu in the porewater.

Figure 4-1. Change in DTPA extractable Cu over time for each treatment: a) CuO NP treatment, b)

CuSO4 treatment, and comparison of mean of extractable Cu for each Cu treatments at the end of

the plant growth period: c) DTPA extraction, d) CaCl2 extraction. Error bars show ± 1 standard

deviation. In a and b, capital letters indicate significant differences between DTPA extractable Cu at

four time points for CuO NP treatments (a) and CuSO4 treatments (b). In c and d, capital letters

indicate significant differences in DTPA extractable Cu (c) and CaCl2 extractable Cu (d) among soils

collected after plant harvesting (ANOVA test followed by Fisher’s LSD test for multiple

comparisons, P≤0.05).

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4.3.3 Toxicity of CuSO4 and CuO NP.

Root maximal length, root compactness (root mass/root maximal length), leaf lengths, shoot mass

(Figure 4-2) and root morphology (Appendix 3, Figure A3-5 and Figure A3-6) were used to evaluate

the toxic effect of CuSO4 and CuO NPs.

Root maximal length and root compactness indicated no visual toxic effect from the fresh

CuO NP treatment. For other treatments, significant decreases in root maximal length (a decrease of

6.6cm, 8.2cm, and 6.8cm compared to the control treatment for aged CuO NPs, fresh CuSO4, and

aged CuSO4, respectively) were observed. Increased root compactness was observed for the aged

CuO NP treatment (an increase of 4.0 mg/cm compared to the control) and for the fresh CuSO4

treatment (an increase of 5.1 mg/cm compared to the control). Examples of shortened roots and

compactness of roots are shown in the appendix (Figure A3-5, Appendix 3). Evidence of Cu toxicity

was also observed in Cytoviva images. In comparison to the roots exposed to CuSO4 (Appendix 3,

Figure A3-6), the roots exposed to CuO NP (fresh or aged) did not present the same damaged

physiology. Roots exposed to CuSO4 (both fresh and aged) showed a brown damaged (necrotic)

zone that was not found on any of the CuO NP exposed roots. No effects of Cu on the shoots (leaf

length, biomass) were observed for the CuO NP treatments. Both the freshly amended and aged

CuSO4 treatments resulted in shorter third leaves (shortened by 5.4cm and 4.0 cm compared to the

control for fresh and aged CuSO4 treatments, respectively). The freshly amended CuSO4 treatment

also had less total shoot biomass compared to the control treatment.

Some indication of toxicity was evident in all treatments except for the fresh CuO NP

treatment. Aging of CuSO4 decreased its toxicity to Triticum aestivum, while aging of CuO NP

increased its toxicity. Overall, the two CuSO4 treatments showed more toxic effects to Triticum

aestivum compared with the two CuO NP treatments, even though the CuSO4 was added at a

significantly lower Cu concentration (300mg/kg for CuSO4 treatments vs. 500mg/kg for CuO NP

treatment).

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Figure 4-2. a) Root compactness and b) leaf length (leaf growth stage is noted with number, from 1

being the youngest to 3 the oldest) of wheat seedlings grown in freshly amended and aged CuO NP,

CuSO4-amended soil, and control treatments. Error bars show ±1SD, * indicates P≤0.05; **

indicates P≤0.01. (ANOVA test followed by Fisher’s LSD test for multipal comparisions) compared

to the control treatment.

4.3.4 Cu root association and Cu uptake

The presence of CuO NPs associated with the roots after 2 weeks of plant growth in both fresh and

aged CuO NP amended soils was investigated using enhanced dark-field hyperspectral imaging (DF-

HSI) as shown in Figure 4-3. The pixels containing CuO NP have been highlighted in red. In both

fresh and aged CuO NP amended soils (Figure 4-3), CuO NPs were found associated to specific

locations on the roots, either to the root tip mucilage (Figure3 a, b, f, g), or to soil aggregates

attached to the root hairs or root tips (Figure 4-3 a, c-i). For the concentration of Cu in roots, all Cu

treatments were significantly higher than the control treatment. The Cu concentration in roots

(577mg/kg, s.d.=46mg/kg, 6 replicates) was statistically significantly higher in the freshly amended

CuO NP treatment than in the aged CuO NP treatment (400mg/kg, s.d.=60mg/kg, 6 replicates) or

either ionic treatment (278mg/kg, s.d.=51mg/kg, 6 replicates for fresh CuSO4 and 442mg/kg,

s.d.=67mg/kg, 6 replicates for aged CuSO4) (Appendix 3, Figure A3-7, a). For the shoot

concentrations, no statistically significant differences were found for all Cu treatments (53mg/kg-

88mg/kg) (Appendix 3, Figure A3-7, b).

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Figure 4-3. Hyperspectral imaging of plant roots grown in soil with freshly amended CuO NPs (a-e)

or after aging (f-i). The b, c and g views are magnified views from a and f. Pixels containing the

reflectance spectra specific to CuO NPs are highlighted in red. CuO NPs and their aggregates were

found associated to mucilage, root tissues and root hairs (red arrow), and to soil aggregates attached

to those locations (yellow arrows). Scale bars: 25µm.

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4.3.5 Effect of near-root environment on Cu availability from CuO NP treatment Figure 4-4 shows the differences in extractable Cu in rhizosphere soil, loosely attached soil and bulk

soil for fresh and aged CuO NP treatments. For the CaCl2 extraction in both fresh and aged CuO

NP treatments (Figure 4-4a, b), the extractability of Cu in the rhizosphere soil was significantly

higher than the extractability of Cu in the loosely attached soil or bulk soil (ANOVA test, P≤0.05).

There were no statistically significant differences among DTPA extractable Cu measurements from

rhizosphere soil, loosely attached soil and bulk soil in the freshly amended CuO NP treatment

(Figure 4-4c). However, the DTPA extractable Cu in the rhizosphere soil in the aged CuO treatment

was significantly lower than the DTPA extractable Cu in bulk soil, but similar to that measured in

loosely attached soil (Figure 4d). In control experiments (Na2SO4), the CaCl2 extractable Cu was

below the detection limit (0.08mg/kg in soil, 4ug/L for the diluted samples) in all soil samples.

Aging increased the concentration of CaCl2 extractable Cu and DTPA extractable Cu in loosely

attached soil and bulk soil, and increased the concentration of DTPA extractable Cu in rhizosphere

soil (t test, P<0.05). But aging did not change the CaCl2 extractable Cu in rhizosphere soil (t test,

P>0.05).

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Figure 4-4. CaCl2 and DTPA extractable Cu in fresh (left side) and aged (right side) CuO NP

amended rhizosphere soil, loosely attached soil and bulk soils. Error bars show ± 1 SD. Capital

letters indicate significant differences between groups (One way ANOVA test followed by Fisher’s

LSD test for multiple comparison, P≤0.05).

4.3.6 Soil pH in bulk soil, rhizosphere soil and loosely attached soil

For all CuO NP treatments and the control treatment (no addition), the pH of the rhizosphere soils

was significantly higher than the pH of the loosely attached and bulk soils (Figure 4-5a, b and c). In

freshly amended CuO NP treatments and control treatments, the pH of the loosely attached soils

were not statistically significantly different than the pH in the bulk soils. However, in aged CuO NP

treatments, the pH of the loosely attached soil was statistically significantly higher than the pH in

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bulk soil. In bulk soil, the pH was the highest in freshly amended CuO NP treatments, followed by

aged CuO NP treatment, followed by the control treatment, followed by the aged CuSO4 treatment,

with freshly amended CuSO4 treatment having the lowest soil pH (Figure 4-5d).

Figure 4-5. Mean ± SD of soil pH (measured using CaCl2 extraction) in rhizosphere soil, loosely

attached soil and bulk soil in a) soil freshly amended with CuO NP, b) aged CuO NP treatment c)

control soil, and; d) Comparison of pH of bulk soil among all treatments. Capital letters indicate

significant differences (ANOVA test followed by Fisher’s LSD test for multiple comparison,

P≤0.05).

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4.4 Discussion

4.4.1 CuO NP dissolution is linked to toxicity.

Compared to Cu ions, the dynamic dissolution process of CuO NP in soil led to a very different Cu

exposure profile for plants. At the end of the two growth periods, the DTPA-extracted Cu

concentrations in CuO NP treatments were similar or even higher than in the CuSO4 treatment.

However, a decreasing trend in DTPA extractable Cu on CuSO4 treatments during the two plant

growth periods was observed. This decrease can be attributed to the soil-organic matter interactions,

solid-state diffusion of Cu ions into iron minerals or metal (co)precipitates52–54. Conversely, an

increase in DTPA-extractable Cu over time was shown in CuO NP treatments (fresh treatment and

aged treatment) during the two plant growth periods and the aging period. This can be attributed to

the dissolution of CuO NP11,12. Thus, the plants in the freshly amended CuO NP soil were exposed

to lower amounts of labile Cu compared to either the two CuSO4 treatments or the aged CuO NP

treatment. These findings suggest that when evaluating the chemical availability or toxic effect of

metal-based NPs in soil, single-time point chemical extractions at the end of the plant growth period

cannot capture the dissolution process of NPs in soil, and thus may fail to predict the toxicity or

bioavailability of NPs11,12,55. Considering that it is not feasible to uniformly dose Cu ions into soil

over time to precisely mimic the dosing rate from NP dissolution, toxicity studies with soluble NPs

should measure the dissolution rate in soil and monitor the behavior of soluble ions in soil, and

interpret their results in light of the different dosing conditions that manifest.

A significantly higher toxicity in CuSO4 treatments compared to the fresh CuO NP

treatment is explained by the higher exposure of roots to labile Cu species, even though the CuO

NP treatment had a higher total Cu concentration. Also, dissolution of CuO NPs over time

gradually increased the available Cu in soil, leading to higher toxicity in the aged treatment. The

opposite has been observed with CuSO4, where the available Cu in soil decreased over time, leading

to lower toxicity to the plants in the aged treatment. The effects of time on toxicity of CuO NP and

CuSO4 have already been observed19. The authors attributed this to the dissolution behavior of CuO

NP, although without quantification. Here, we clearly showed that in order to correlate the chemical

availability of CuO NPs with toxicity, the dissolution kinetics, i.e. predicting the total Cu released to

soil during the growth period, should be considered. The dissolution kinetics can be modeled as

first-order dissolution, with the rate constant fit to the extractable Cu over time11, and the total

amount of Cu ion released from CuO NPs can be estimated by integrating the expression relating

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the change of extractable Cu over time. This observation is relevant to NP formulations of

fungicides and micronutrients, so the release rate of the active ingredients can be better timed to the

plant’s needs.

4.4.2 CaCl2 extractable Cu correlates with toxicity of CuO NP to wheat.

Although DTPA extractable Cu gives a better indication of CuO NP dissolution because it extracts

most of the Cu species in soil, CaCl2 extractable Cu is better for correlating toxicity, since it

measures dissolved Cu in pore water that can directly interact with plant roots. DTPA extraction

would predict the toxicity of the aged CuO NP to be higher than the CuSO4 treatment (Figure 4-1 a,

b). However, this was not the case. The aged CuO NP had lower toxicity compared to both the aged

and the fresh CuSO4 treatment, indicating that the DTPA extraction cannot accurately predict

toxicity for the CuO NPs. The CaCl2 extraction ranked them correctly (Figure 1d). Considering that

the extractable Cu in CuO NP amended soil increased over time while the extractable Cu in CuSO4

amended soil decreased over time in Lufa 2.2 soil (Appendix 3, Figure A3-1), the wheat plants were

exposed to a lower overall ‘readily available’ Cu (i.e. Concentration x time) in CuO NP treatments

compared to the CuSO4 treatments. The lower CaCl2 extractable Cu in aged CuO NP treatment is a

result of higher soil pH in aged CuO NP treatment compared to the fresh and aged CuSO4

treatment. Higher soil pH has been previously shown to lower Cu concentration in soil pore water 27,56.

4.4.3 Root-associated CuO NP modulates toxicity. In the freshly amended CuO NP soil, although being exposed to a lower concentration of labile Cu,

the roots of Triticum aestivum were actually exposed to higher total Cu concentration (Appendix3,

Figure A3-7) than the other treatments. This is mainly due to CuO NPs' association with plant roots

(Figure 4-3b). This exposure to CuO NPs did not lead to any detected toxic effects, indicating a low

or de minimis level of toxic effects from the particle itself over the 14d growth period.

4.4.4 Root exudates affect CuO NP dissolution and availability. The increase in the pH of rhizosphere soil compared to bulk soil in our study indicates that the

rhizosphere region was indeed influenced by the plant roots. Excretion of organic acid (dissociated

ions) by plant roots , nitrogen uptake and ionic exchanges by plant roots may explain the higher pH

of the rhizosphere soil compared to the pH of bulk soi39,40,57,58. The observed pH change in

rhizosphere soil was not likely a result from the presence of CuO NP, as similar pH changes

occurred with both the fresh CuO NP treatment and the negative control treatment (0.4pH unit,

- 69 -

ANOVA test, P>0.05). However, in the aged CuO NP treatment, the pH increase was higher

(0.6pH unit, ANOVA test, P<0.05 ), potentially because the release of Cu2+ from CuO NP attached

to the roots triggered more root responses. A previous study has shown that a high concentration of

ionic Cu can increase root exudation59. The other factor that may affect CuO NP dissolution besides

pH is the organic acids released by plants. Although we did not measure them here, their release is a

well-known mechanism by which Triticum aestivum increases the availability of nutrients and

decreases the toxic effects of metal ions such as Cu and Al36–38.

Influences of rhizosphere soil pH and root exudates on the dissolution of the CuO NPs and on the

lability of Cu derived from CuO NP were revealed by the two extraction methods (Figure 4-4).

Though rhizosphere soil exhibited higher pH (which should reduce the concentration of free Cu

ions), more Cu was extracted by CaCl2 from rhizosphere soil relative to bulk soil, indicating a greater

amount of complexed Cu ions in rhizosphere soil pore water. This complexation is likely a result of

small organic acids released by plants and their microbiomes in the rhizosphere region. The

influence of the roots on the extent of dissolution of CuO NPs was shown by the extraction of

“labile” Cu (DTPA). For the aged CuO NP treatment, the DTPA extractable Cu in rhizosphere soil

was lower than for the bulk soil. DTPA extraction has been shown to extract most of the labile Cu

species in soil (~80%), but it cannot extract CuO NP11(Appendix 3, Figure A3-1). Thus, a reduction

in DTPA extractable Cu suggests diminished CuO NP dissolution in rhizosphere soil. This is

consistent with the slightly higher measured pH in rhizosphere soil compared to bulk soil (Figure 4-

5), which decreases the dissolution rate and solubility of CuO NP. 10,12 This difference is less evident

in the fresh CuO NP treatment than the aged CuO NP treatment, consistent with the slightly lower

rhizosphere soil pH in the fresh CuO NP treatment compared to the aged CuO NP treatment

(Figure 4-5).

For CuO NP treatment, the DTPA extractable Cu in all soil compartments (rhizosphere soil, loosely

attached soil and bulk soil) increased from the fresh treatment to the aged treatment as a result of

CuO NP dissolution. This dissolution also resulted in higher ‘readily available’ CaCl2 extractable in

loosely attached soil and bulk soil. However, in rhizosphere soil, the impact of CuO NP dissolution

on the ‘readily available’ Cu was less significant, possibly because complexation by root exudates

played a more important role in increasing the ‘readily available’ Cu in rhizosphere soil.

The interaction between plant, soil, and Cu was limited to the rhizosphere soil region during

the 14-d plant growth period, as suggested from the measured pH and extractable Cu in different

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soil zones. This finding is consistent with a previous study showing the pH was affected by durum

wheat roots only within a few millimeters60. Even the loosely attached soil collected in this study

remained mostly unaffected by the presence of plants. However, the spatial extent where root

exudates may affect the dissolution of CuO NPs over longer growth periods needs more

investigation. The changes in soil pH and the release of plant exudates changed the chemical

availability and the dissolution rate of CuO NPs in rhizosphere soil. Thus, measuring the changes of

CaCl2 extractable Cu over time in rhizosphere soil may be the best way to correlate CuO NP

dissolution with its toxicity. Further study is needed to quantitatively evaluate the impact of root

exudates and changes in pH on dissolution behavior of CuO NP in rhizosphere soil over time to

decide whether rhizosphere soil should routinely be used in extraction tests to predict the toxicity of

CuO NPs to terrestrial plants. Bacteria in root rhizosphere may also play a role in CuO NP

availability, dissolution and uptake via exuding chelating compounds like pyoverdines (siderophores

produced by certain pseudomonads that can complex with metal ions). Future studies needs to

elucidate the role of these bacteria have in rhizosphere region61,62. However, these experiments

indicate that measurements of Cu availability in bulk soils will not likely provide an accurate

representation of the Cu availability to the plants.

4.4.5 Triticum aestivum regulated Cu uptake. Although we have shown a strong correlation between CuO NP dissolution and the toxicity of

CuO NP to Triticum aestivum, there was no such correlation between CuO NP dissolution and Cu

translocation in Triticum aestivum shoots. Despite being exposed to different concentrations of total

and labile Cu species during the growth periods among all treatments, as shown in Figures 1a and b,

Triticum aestivum tended to take up similar amounts of Cu (Appendix 3, Figure A3-7). This suggests

that the uptake of Cu was regulated by Triticum aestivum, consistent with a previous study that found

low Cu uptake by wheat even when the labile Cu concentration in soil was high63. When Cu uptake

is regulated by a plant, the bioavailable Cu in soil measured by extraction cannot be correlated to Cu

uptake.

4.5 Agricultural implications Our study indicates some potential benefits of using nano-CuO as a micronutrient amendment or

fungicide rather than the soluble Cu salt. Firstly, CuO NPs were less toxic than CuSO4, despite being

applied at a much higher dose to the soils. The slow dissolution of CuO NPs reduced the maximum

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concentration of CaCl2-extractable Cu experienced by plants, providing a continuous release of Cu

over 14 days without showing any visual toxic effect. After aging for 28d, some Cu phytotoxic

effects were observed. However, the dissolution rate of a ‘nano-enabled’ fertilizer or fungicide could

likely be tuned to provide sustained release the Cu ions at a rate where the concentration of Cu ions

would not exceed the phytotoxic concentration22. This tuning could potentially be accomplished

through surface modification, addition of other phases (e.g. via doping), using mixtures of different

sized particles, or adjusting their size or aspect ratio.

Secondly, we have shown that CuO NPs have higher affinity to plant roots than Cu ions. Thus,

‘targeting’ the NPs to plant roots could be another potential benefit for nano-enabled fertilizer. In a

calcareous soil with high pH, the dissolution of CuO NP in bulk soil would be very low, but plant

exudates could still potentially enhance the dissolution behavior of CuO NP and the availability of

Cu in the rhizosphere soil, avoiding the toxicity of excess of Cu. This may also reduce Cu

accumulation in soil, a problem with many Cu-containing fertilizers/pesticides. The improvements

in nutrient uptake efficiency or antifungal properties due to this ‘targeting’ requires more

investigation.

Conceivably, nano-enabled, slow-release fertilizers could thus be designed to be applied at high

concentration but low toxicity, to last for years without re-application. This would save energy and

labor, incentives to growers to adopt those new technologies. Research is still needed to determine

the rate of delivery that can provide its function effectively, but without invoking toxicity.

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(46) Duffner, A.; Weng, L.; Hoffland, E.; van der Zee, S. E. Multi-surface modeling to predict free zinc ion concentrations in low-zinc soils. Env. Sci Technol 2014, 48 (10), 5700–5708.

(47) Jeong, C. Y.; Park, C. W.; Kim, J.-G.; Lim, S. K. Carboxylic content of humic acid determined by modeling, calcium acetate, and precipitation methods. Soil Sci. Soc. Am. J. 2007, 71 (1), 86–

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94.

(48) Ritchie, J. D.; Perdue, E. M. Proton-binding study of standard and reference fulvic acids, humic acids, and natural organic matter. Geochim. Cosmochim. Acta 2003, 67 (1), 85–96.

(49) Villaverde, P.; Gondar, D.; Antelo, J.; Lopez, R.; Fiol, S.; Arce, F. Influence of pH on copper, lead and cadmium binding by an ombrotrophic peat. Eur. J. Soil Sci. 2009, 60 (3), 377–385.

(50) 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 (10), 375–380.


(52) Dale, A. L.; Casman, E. A.; Lowry, G. V; Lead, J. R.; Viparelli, E.; Baalousha, M. Modeling nanomaterial environmental fate in aquatic systems. Env. Sci Technol 2015, 49 (5), 2587–2593.

(53) Utembe, W.; Potgieter, K.; Stefaniak, A. B.; Gulumian, M. Dissolution and biodurability: Important parameters needed for risk assessment of nanomaterials. Part. Fibre Toxicol. 2015, 12 (1), 11.

(54) McGivney, E.; Gao, X.; Liu, Y.; Lowry, G. V; Casman, E. A.; Gregory, K. B.; Vanbriesen, J. M.; Avellan, A. Biogenic cyanide production promotes dissolution of gold nanoparticles in soil. Environ. Sci. Technol. 2018.

(55) Avellan, A.; Simonin, M.; McGivney, E.; Bossa, N.; Spielman-Sun, E.; Rocca, J. D.; Bernhardt, E. S.; Geitner, N. K.; Unrine, J. M.; Wiesner, M. R. Gold nanoparticle biodissolution by a freshwater macrophyte and its associated microbiome. Nat. Nanotechnol. 2018, 13 (11), 1072.

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CHAPTER 5: Dissolution functional assay improves understanding of metallic nanoparticle toxicity in agricultural soil

Abstract Cu-based ENMs have been used in agriculture. Functional assays are needed to predict their

environmental risks. The objective of this study was to use the dissolution profiles of Cu-based

agrochemicals in soil as a functional assay to predict the toxicity of Cu-based ENMs to Triticum

aestivum. Dissolution kinetics of different Cu-based ENMs, including CuO NP and Cu(OH)2 NP

were measured in standard 2.2 soil with a diethylenetriamine pentaacetic acid (DTPA) extraction

method. DTPA-extractable Cu (i.e. soluble Cu species) was plotted over time to get the dissolution

profile. Biological end points, including the toxicity of Cu-based ENMs to Triticum aestivum (reflected

by reduced maximum root length). The integrated exposure to total soluble Cu (the area under the

dissolution curve corresponding to the exposure interval) was correlated with the biological

endpoint. The integrated exposure to total soluble Cu in different Cu species treatments correlated

well with its toxicity to Triticum aestivum (R2=0.91). In contrast, the standard measure of exposure

(DTPA extractable Cu measured at the end of exposure period) failed to correlate with observed

toxicity (R2=0.09). This study suggested that the integrated exposure of Cu-based ENMs may be a

better measure of exposure than DTPA extractable Cu at the end of an experiment.

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5.1 Introduction Pesticidal, micronutrient, or fungicidal copper-based engineered nanomaterials (ENMs)

could enter agricultural soil though either a direct soil application or indirectly, as overshoot from

foliar application of agrochemicals1–3. It is thus important to understand the fate and bioavailability

of Cu-based ENMs in soil to determine the appropriate application rate of the material, and to

evaluate the risk of the ENMs to non-targeted organisms.

Previous studies have shown that ENMs exhibited a different toxicity than macroparticles or soluble

ions. In aqueous systems, the toxicity of ENMs is usually attributed to the dissolution of ENMs that

releases the soluble ions4–7. Different mechanisms have been proposed to explain the toxicity of

ENMs in soil, either attributing toxicity to a nano-specific effect8 or to ion release9–11. Different

studies have concluded that the same ENMs exerted their toxicity exclusively through one or the

other mechanism. Differences in soil characteristics account for some of the confusion12. This paper

clarifies the seemingly contradictory conclusions of toxicological studies and demonstrates how,

with appropriate functional assays1 to measure the dissolution profile of ENMs in these media,

much of this inconsistency can be explained. Dissolution profile is defined in this paper as the area

under a dissolution curve (dissolved metal concentration over time), as suggested by the shaded

region in Figure 5-1. The objective of this study was to investigate how dissolution profiles

measured directly in soil can explain Cu-based ENMs’ toxicity to wheat plants in soils.

5.2 Methods 5.2.1 Dissolution profile measurement assay

Cu-based ENMs were added to soil and well mixed with wooden sticks. At different aging

periods (T=0h, 1d, 2d, 7d, 14d and 28d, plus T=5h for particles with fast dissolution kinetics), 2.0 g

of air-dried soil were extracted with 4 ml of 0.005M DTPA and 0.1M triethanolamine (TEA)

(pH=7.6) to measure the total soluble Cu released from Cu-based ENMs due to dissolution. Sample

1 Functional assays are semi-empirical methods to measure the processes or functions in a

particular system that can relate to certain endpoints, e.g. toxicity and bioaccumulation19.

Importantly, the outcomes of functional assays are determined by both the properties of the

materials tested and the properties of the environmental systems.

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bottles were horizontally shaken for 2 hours for the extraction. After extraction, all samples were

centrifuged at 3000 rpm for 10 min, and the supernatants were filtered using a 0.2 µm PTFE filter.

All samples were acidified with 20% HNO3 in Milli-Q-water (final HNO3 concentration was 2%)

and analyzed by ICP-MS (Agilent 7700).

DTPA extractable Cu was then plotted over time to get the profile of the dissolution curve,

which was integrated to produce a “integrated exposure to total soluble Cu”.

5.2.2 Plant uptake measurement and toxicity The details of the experimental design is discussed in Chapter 413. Briefly, Triticum aestivum

seedlings were harvested after being exposed to CuO NPs for 14 days in Lufa 2.2 soil. Two

treatments for each copper species were tested, the first when the wheat seedling was transferred

into the copper-amended soil at day 0 and the second when transplantation began on day 28 after

ENM amendment. The latter simulates the aging process in soil. Plant roots and shoots were

photographed with a scale bar for determination of length, as an indication of plant health. In this

study, root length was used as the biological endpoint, as it was the most sensitive toxicity effect

observed for Triticum aestivum.

5.3 Results and discussion:

5.3.1 Differences in dissolution time scale require different assays In soils, different Cu species would have different dissolution profiles (Figure 5-1). For CuO

NP (Figure 5-1 a), the concentration of total soluble Cu in bulk soil increases relatively slowly over

time, before plateauing at this particle’s solubility. In contrast, Cu(OH)2 NPs (Figure 5-1 b)

undergoes an initial fast dissolution followed by sorption to soil organic matter that reduces Cu2+

concentration in soil. For the dissolved species, CuSO4 (Figure 5-1 c), the Cu2+ concentration

remained constant over time after a brief initial equilibration period. These different dissolution

profiles suggest organisms exposed to the same initial doses of different Cu soil amendments will be

exposed to different amounts of Cu2+ initially and over time, and that measuring toxicity without

regard to the dissolution profile will result in noisy at best, and incoherent at worst, results.

Functional assays that measure the actual metal ion exposures therefore should be a part of

nanoparticle toxicity testing protocols in soils.

When the dissolution process is essentially complete before the beginning of the exposure

period, (e.g. CuSO4, Figure 5-1, c ), a single extraction assay measuring the available Cu at the end of

the exposure period will be sufficient.15,16 However, for slowly dissolving NPs (Figure 5-1, a) , or if

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the dissolution profile is complex (Figure 5-1, b,), a single extraction at the end of the experiment is

inadequate because no single time point represents the whole exposure. Instead, integration of the

dissolution curve over the time of exposure to get the integrated exposure is recommended.

0 1 0 2 0 3 0 4 0

0

5 0

1 0 0

1 5 0

2 0 0

T i m e ( d a y s )

Ex

tra

cta

ble

Cu

(mg

/k

g d

rie

d s

oil

)

C u O N P s , L u f a 2 . 2 s o i la )

0 1 0 2 0 3 0 4 0

0

1 0 0

2 0 0

3 0 0

C u ( O H ) 2 , L u f a 2 . 2 s o i l

T i m e ( d a y s )

Ex

tra

cta

ble

Cu

(mg

/k

g d

rie

d s

oil

)

b )

0 1 0 2 0 3 0 4 0

0

1 0 0

2 0 0

3 0 0

T i m e ( d a y s )

Ex

tra

cta

ble

Cu

(mg

/k

g d

rie

d s

oil

)

C u S O 4 , L u f a 2 . 2 s o i lc )

Figure 5-1. The dissolution profile of 250mg/kg of CuO NP, Cu(OH)2 NP and CuSO4 in Lufa 2.2 soil.

5.3.2 Dissolution profile measurement assay predicted toxicity of Cu species to Triticum aestivum Cu-based ENMs in soil usually demonstrated a much slower and more complex Cu release profile,

than soluble Cu species. A solubility measurement at one time point is not sufficient to capture the

different dissolution profiles in soil13,17. To correlate the biological endpoints of Cu-based ENMs

applied to soil (in this study, toxicity), a dissolution profile is needed.

When Triticum aestivum were exposed to fresh and aged CuO NPs and CuSO413, their toxicities

(shortened root length) to Triticum aestivum were correlated with the integrated exposure to total

soluble Cu (arena under the dissolution curve, R2 = 0.91, Figure 5-2 (a)). On the other hand, the

traditional way of measuring bioavailability, which is, using the DTPA extractable Cu at end of the

exposure period, failed to predict the toxicity of Cu species to wheat plants ( R2 = 0.09, Figure 5-2

(b)).

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0 1 0 0 0 2 0 0 0 3 0 0 0 4 0 0 0 5 0 0 0

0

2

4

6

8

1 0

I n t e g r a t e d e x p o s u r e t o t a t a l s o l u b l e C u

( m g * d a y / k g )

Sh

ort

en

ed

ro

ot

ma

xim

al

len

gth

(c

m)

C u O N P ( f r e s h )

C u O N P ( a g e d )

C u S O 4 ( a g e d )

C u S O 4 ( f r e s h )

R2

= 0 . 9 1

( a )

2 0 0 2 5 0 3 0 0 3 5 0

0

2

4

6

8

1 0

C o n c e n t r a t i o n a t t h e e n d o f e x p o s u r e p e r i o d ( m g / K g )Sh

ort

en

ed

ro

ot

ma

xim

al

len

gth

(c

m)

C u O N P ( f r e s h )

C u S O 4 ( a g e d )

C u S O 4 ( f r e s h )

C u O N P ( a g e d )

R2

= 0 . 0 9

( b )

Figure 5-2. Correlations between (a) integrated exposure to total (DTPA-extractable) soluble Cu

and toxicity to Triticum aestivum (shortened root maximal length) and (b) Cu2+ concentration at the

end of exposure period (28 or 42 days of aging) and toxicity to Triticum aestivum (shortened root

maximal length)13. The word “fresh” indicates that plants were exposed to Cu-amended soil right

after the Cu species were applied to soil. “Aged” means the plants were exposed to Cu amended soil

after 28-day aging periods.

The data presented, though suggestive of an important governing principle for soil nanotoxicity,

only cover 2 Cu species (CuO NP and CuSO4) and one plant (Triticum aestivum).In order to be

convincing, this principle would have to be demonstrated for more nanoparticles and more

organisms.

5.3 Environmental Implications

This study suggested that Cu-based ENMs in soil usually require a complete characterization

of the dissolution profile to capture ionic exposure, as a one time measurement of Cu ion

availability is not sufficient to capture the different dissolution profiles 13,17. Previous studies have

developed models that can predict the dissolution profile of CuO NP by knowing the soil pH and

organic carbon content18, which will potentially enable us to predict exposure without measuring

dissolution. In this case, one of the future challenges would be to extend such models and the

functional assays on which they are based, to other types of particles.

This study also raises a concern about the traditional definition of the chronic toxicity test for

nanoparticles in soil. Due to the dissolution behavior of the ENMs, they transformed during a

toxicity test. This suggests that the test result from the toxicity test may not reflect the risk and

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toxicity of the pristine material, but the transformed material (e.g. the effect might be caused by the

released Cu2+ from CuO NP, not the pristine CuO NP). Thus, future studies need to reconsider the

chronic toxicity test for soluble ENMs in soil, with the consideration of their dissolution profile. In

this case, the concept of dissolution profile discussed in this study could be used to better reflect the

exposure of the material instead of concentration. Though not explored here, the variety of shapes

of dissolution profiles may have implications for acute soil nanotoxicity testing in soil as well. Should

the acute exposure coincide with the peak dissolution or be measured always for the same duration

(without consideration of the shape of dissolution profile)?

5.4 References for Chapter 5

(1) Tegenaw, A.; Tolaymat, T.; Al-Abed, S.; El Badawy, A.; Luxton, T.; Sorial, G.; Genaidy, A. Characterization and potential environmental implications of select Cu-based fungicides and bactericides employed in U.S. markets. Env. Sci Technol 2015, 49 (3), 1294–1302.

(2) Rodrigues, S. M.; Demokritou, P.; Dokoozlian, N.; Hendren, C. O.; Karn, B.; Mauter, M. S.; Sadik, O. A.; Safarpour, M.; Unrine, J. M.; Viers, J. Nanotechnology for sustainable food production: promising opportunities and scientific challenges. Environ. Sci. Nano 2017, 4 (4), 767–781.

(3) Peters, R. J. B.; Bouwmeester, H.; Gottardo, S.; Amenta, V.; Arena, M.; Brandhoff, P.; Marvin, H. J. P.; Mech, A.; Moniz, F. B.; Pesudo, L. Q.; et al. Nanomaterials for products and application in agriculture, feed and food. Trends Food Sci. Technol. 2016, 54, 155–164.

(4) Blinova, I.; Ivask, A.; Heinlaan, M.; Mortimer, M.; Kahru, A. Ecotoxicity of nanoparticles of CuO and ZnO in natural water. Environ. Pollut. 2010, 158 (1), 41–47.

(5) Starnes, D. L.; Unrine, J. M.; Starnes, C. P.; Collin, B. E.; Oostveen, E. K.; Ma, R.; Lowry, G. V; Bertsch, P. M.; Tsyusko, O. V. Impact of sulfidation on the bioavailability and toxicity of silver nanoparticles to Caenorhabditis elegans. Env. Pollut 2015, 196, 239–246.

(6) Xiu, Z.; Zhang, Q.; Puppala, H. L.; Colvin, V. L.; Alvarez, P. J. J. Negligible particle-specific antibacterial activity of silver nanoparticles. Nano Lett. 2012, 12 (8), 4271–4275.

(7) Levard, C.; Hotze, E. M.; Colman, B. P.; Dale, A. L.; Truong, L.; Yang, X. Y.; Bone, A. J.; Brown Jr, G. E.; Tanguay, R. L.; Di Giulio, R. T. Sulfidation of silver nanoparticles: natural antidote to their toxicity. Environ. Sci. Technol. 2013, 47 (23), 13440–13448.

(8) Servin, A. D.; Pagano, L.; Castillo-Michel, H.; De la Torre-Roche, R.; Hawthorne, J.; Hernandez-Viezcas, J. A.; Loredo-Portales, R.; Majumdar, S.; Gardea-Torresday, J.; Dhankher, O. P. Weathering in soil increases nanoparticle CuO bioaccumulation within a terrestrial food chain. Nanotoxicology 2017, 11 (1), 98–111.

(9) Watson, J.-L.; Fang, T.; Dimkpa, C. O.; Britt, D. W.; McLean, J. E.; Jacobson, A.; Anderson,

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A. J. The phytotoxicity of ZnO nanoparticles on wheat varies with soil properties. Biometals 2015, 28 (1), 101–112.

(10) Dimkpa, C. O.; Hansen, T.; Stewart, J.; McLean, J. E.; Britt, D. W.; Anderson, A. J. ZnO nanoparticles and root colonization by a beneficial pseudomonad influence essential metal responses in bean (Phaseolus vulgaris). Nanotoxicology 2015, 9 (3), 271–278.


(12) Neves, J.; Cardoso, D. N.; Malheiro, C.; Kah, M.; Soares, A. M. V. M.; Wrona, F. J.; Loureiro, S. Copper toxicity to Folsomia candida in different soils: a comparison between nano and conventional formulations. Environ. Chem.


(14) Standardization, I. O. for. Soil Quality-inhibition of Reproduction of Collembola (folsomia Candida) by Soil Pollutants; International Organization for Standardization, 1999.

(15) Chojnacka, K.; Chojnacki, A.; Gorecka, H.; Gorecki, H. Bioavailability of heavy metals from polluted soils to plants. Sci Total Env. 2005, 337 (1–3), 175–182.

(16) Sillanpää, M.; Oikari, A. Assessing the impact of complexation by EDTA and DTPA on heavy metal toxicity using microtox bioassay. Chemosphere 1996, 32 (8), 1485–1497.


(18) Gao, X.; Rodrigues, S. M.; Spielman-Sun, E.; Lopes, S. P.; Rodrigues, S.; Zhang, Y.; Avellan, A.; Duarte, R. M. B. O.; Duarte, A. C.; Casman, E. A. Effect of soil organic matter, soil pH, and moisture content on solubility and dissolution rate of CuO NPs in soil. Environ. Sci. Technol. 2019.

(19) 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.

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CHAPTER 6: Summary of Major Contributions and Perspective on Future Research

6.1 Summary of Major Contribution This dissertation has made contributions towards the understanding the behavior of engineered

nanomaterials (ENMs) in soil systems, and their impact on soil organisms. The novelty as well as

contributions of this thesis are summarized here.

6.1.1. Major Contribution from Objective 1: A test method to measure dissolution of CuO NP in soil was developed. Two time-based chemical extraction methods were developed to measure the dissolution kinetics of

CuO NPs directly in a soil matrix. The first method is based on a pore-water extraction method1,2

(CaCl2 extraction) that measures the dissolution by monitoring the concentration of released Cu2+

from CuO NP in soil pore water. The second method is based on a labile metal extraction method3

(DTPA extraction) that measures the dissolution by monitoring the reversibly-sorbed fraction

(~80%) of the Cu released from CuO NP4. Both extractions were performed at different aging times

after CuO NPs were dosed to soil. The increase in extractable Cu was fit with a kinetic model to

quantify the dissolution rate constant. This study provided researchers in this field a tool to evaluate

the dissolution of Cu-based ENMs in soil systems.

6.1.2. Major Contribution from Objective 2: A model was developed to evaluate the effect of soil pH and organic carbon content on dissolution kinetics of CuO NP in soil. It is unclear from previous studies how soil properties affect the dissolution kinetics of CuO NP in

soil. In this project, we used the tool developed from Objective 1 to measure the dissolution kinetics

of CuO NP in various soils to answer the fundamental question: How do soil properties affect the

dissolution kinetics of CuO NP in soil? Soil organic matter (SOM) content showed a positive

correlation with the solubility of CuO NP in soil, but did not affect the overall dissolution rate

constant (half-life) of the CuO NPs in soil, whereas soil pH affected the dissolution rate constant of

the dissolution process but showed no effect on the ultimate solubility of CuO NP in soil. Soil

moisture content had minimal effect on dissolution kinetics of CuO NP in soil. Based on these

observations, an empirical model to correlate the CuO NP dissolution kinetics with both SOM

content and soil pH was developed. This dissolution model successfully predicted the dissolution

kinetics of CuO NPs in two unknown soil. This work provides fundamental insights on why SOM

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and soil pH play such an important role on the dissolution kinetics of CuO NP in soil. The

dissolution model could also help to explain the differences in behavior (e.g. toxicity) that has been

observed in different soils. The method that we used to develop the dissolution model for CuO NP

could potentially be extended to other metal/metal oxide particles with low to moderate dissolution

rates (half-lives of days to years).

6.1.3. Major Contributions from Objective 3:Dissolution of CuO NPs in soil was correlated with its toxicity to wheat (Triticum aestivum). Dissolution of CuO NPs under the influence of root activity in rhizosphere soil was quantified. Although previous studies have postulated that dissolution of ENMs would affect its toxicity in

soil5–7, the correlation between ENMs dissolution and their toxicity to soil organisms could not be

made without measurement of their dissolution in soil. Using the dissolution measurement method

developed from Objective 1, this work directly correlated the dissolution of CuO NP with its

toxicity to wheat plants for the first time. The increase in toxicity of CuO NP after aging in soil

relative to unaged CuO NPs is explained by CuO NP’s slow dissolution behavior. Here, the impact

of plant roots on dissolution and lability of CuO NP in rhizosphere soil were also quantified for the

first time. This study emphasized the role of ENMs dissolution in understanding their toxicity,

including properly identifying any “nanoparticle specific” toxicity effects.

6.1.4. Major Contribution from Objective 4: Dissolution kinetics functional assays were used to estimate exposure to ionic Cu from Cu-based ENMs in soil. This exposure correlated to observed toxicity in wheat. With the method developed in Objective 1 to measure the dissolution kinetics of CuO NP in soil,

and the observed correlations between dissolution of CuO NPs and its toxicity to wheat in

Objective 3, this work further extended the understanding of the relationship between ENM

dissolution in soil and their toxicity. Here, a functional assay to measure the dissolution profile, i.e.

available Cu concentration over time, of Cu based ENMs (including CuO NP and Cu(OH)2 NP)

was used to predict exposure to Cu ion over time. This exposure correlated to toxicity to wheat

roots demonstrating the importance of properly quantifying exposures over time when dealing with

ENMs in soil.

Overall, this thesis illustrated a way to study dissolution behavior of ENMs in soil. Aging time was

shown to be one important factor to be considered when investigating dissolution. The dissolution

model for CuO NP in soil provided mechanistic understanding on how soil properties affect the

dissolution of ENMs in aerobic soil. It should be noted that this model is based on lab conditions in

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natural but standard soils, not field conditions, which are more complex. For example, soils may

become anerobic for a period of time which could affect the speciation of Cu and the resulting CuO

NP behavior. However, the behavior of CuO NPs in aerobic field soils can be predicted with the

dissolution model developed in this thesis. Leaching is an important process to consider under field

conditions, it essentially removes mobile Cu2+ from the soil. According to the dissolution model

developed in this work, leaching would likely only remove a minor fraction of the Cu in system since

most Cu is sorbed to organic matter and soil solid surfaces rather than mobile in the porewater.

Thus, it may not significantly change the overall dissolution kinetics of CuO NP after all, depending

on the duration of the leaching.

This thesis will be of interest to nanotoxicologists because the measurement of dissolution of ENMs

in soil could help them understand the mechanism of nanotoxicity and precisely define NP exposure

in standard testing. By knowing the dissolution kinetics in soil, the works in this thesis will also guide

the design of nano-enabled agrochemicals. Environmental regulators will also learn from this thesis

that regulations on ENMs in soil might be soil-property-dependent.

6.2 Perspectives for future research Although already being used in agriculture, the risk and environmental impacts of nano-

enabled agrochemicals still need to be evaluated. While it is yet to be known if nano-enabled

agrochemicals will significantly replace the metallic salts traditionally found in agrochemicals, nano-

enabled agrochemicals could be designed to have high efficiency with low environmental risks.

Dissolution is one of the most important processes for metal and metal oxide ENMs in soil, as it

releases the active ingredients, releasing metal ions into the soil. This thesis provides a tool to

measure the dissolution of metal and metal oxide ENMs in soil, as well as models to predict the

dissolution behavior of CuO NP in soils to better quantify soil organisms’ exposure to ions released

from the ENMs. However, there is still more work needed to ensure the efficiency and safety of

nano-enabled agrochemicals. This section will address relevant key challenges that are needed to

guarantee an efficient and safe application of nano-enabled agrochemicals.

6.2.1. Extension of the model to predict the behavior of other metal/metal oxide ENMs in soil CuO NPs are only one type of nano-enabled agrochemical being used in agriculture. Other

metal/metal oxide ENMs, such as ZnO NPs, Ag NPs, Fe2O3 NPs, Fe (OH)3 NPs, Si NPs all have

the potential to be applied in agriculture, either as micronutrient suppliers, or as

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fungicides/pesicides8,9. Also they can enter agricultural soils via accidental release10. While soil pH

and organic matter were considered as the key properties influencing the dissolution kinetics of CuO

NPs, other processes (e.g. redox reactions) will likely be important, as well, for some other NPs and

these need to be addressed in relevant cases. Thus, it could be important to extend the models for

predicting the dissolution kinetics under different redox scenarios. There are a few major challenges

that need to be accomplished to meet such a goal.

First, the Cu speciation change in soil over time is needed. In Chapter 3, the partitioning

between free Cu2+ and Cu2+ associated with different soil surfaces (SOM, clay, etc.) was simplified

using one partitioning constant, Kd. This method worked for this particular model, as SOM is the

predominant soluble Cu sink in soil11,12. However, this may not be the case for other metals.

Coupling the kinetic dissolution model of CuO NP with the equilibrium Cu speciation model (multi-

surface model) would enable us to get a more precise speciation of Cu in soil over time. Such an

exercise would be important to understand the detailed metal speciation in soil dosed with ENMs as

a function of time.

Second, in this thesis, the effect of particle properties on dissolution of CuO NPs in soil was

not investigated. Particle size and shape have been suggested to be important factors affecting the

dissolution kinetics of ENMs in aqueous systems13. The size and shape of ENMs could also

potentially affect their dissolution kinetics in soil, by affecting the ion-release step of the dissolution,

which was suggested in Chapter 3 to be the rate-limiting step. Future work is needed to incorporate

the size and shape properties of CuO NP into the dissolution model.

Third, the extraction methods may need to be modified to measure the dissolution rate of

other metal/metal oxide ENMs in soil. This is required to collect the data needed to build the

model. As summarized in Section 6.1.1, the two extraction methods worked well on measuring the

dissolution of CuO NPs in soils. The success of applying the two chemical extraction methods is

because 1) For CaCl2 extraction, previous studies have shown that the partitioning of Cu2+ between

the soil solid phases and soil pore water is relatively stable in a given soil media11,14. 2) For DTPA

extraction, Cu2+ has a very high stability constant with DTPA15,16, making DTPA a strong chelating

agent capable of extracting most of Cu2+ released from CuO NP in soil. These two necessary

conditions may not be met for other metal/metal oxide ENMs in soil. For example, DTPA does not

chelate strongly with Silver(Ag)17, making it a low efficiency extractant for Ag+ ions from soil. The

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extraction conditions (e.g. extractant used, extraction time, etc.) should be adjusted for different

species of metal/metal oxide ENMs to achieve the best efficiency.

Fourth, the dissolution model needs to be extended to include other important

transformation processes. For example, redox reactions need to be considered for many metals that

are redox active, including Cu2+. The experimental condition applied in Objective 2 was strongly

oxidizing. In agriculture, agrochemicals are usually applied to, and remain in the top soil18, where soil

conditions are generally aerobic. However, there are cases in agriculture that would create anaerobic

conditions, such as paddies of lowland rice, where anaerobic conditions would result from a flood-

irrigation system19, or if the soils flood temporarily. The chemical transformations of many metals

are sensitive to changes in redox potential of the environment20. In anaerobic conditions, metals

such as Cu2+, could be reduced to Cu1+ species or to Cu0, affecting their dissolution behavior. In

fact, a previous study has already observed that CuO NPs were reduced to Cu(I) species in a rice

growing environment21. In this case, redox reactions must be considered when understanding the

dissolution and toxicity of the CuO NPs to rice. To incorporate redox processes into the dissolution

model is necessary to understand the full picture of transformation and dissolution of metal and

metal oxide ENMs in soil/sediment with anaerobic conditions. One potential method for doing this

could be to measure the dissolution rate constant under different redox conditions, and then

correlate dissolution rate constant with redox potential.

Greater understanding of how the composition of SOM and DOM affect the dissolution of

kinetics of CuO NP in soil is also needed. As discussed in Chapter 3, one of the reasons for the

uncertainty in predicting the dissolution profile is the assumption of similar SOM quality (thus same

complexing capacity with Cu) among soils. With more specific characterization of SOM and DOM

composition, it may indeed be possible to improve the predictive capability of the model.

Finally, root exudates need to be better characterized to refine modeling of dissolution of

metal/metal oxide ENMs in the soil rhizosphere. While dissolution of CuO NP in the rhizosphere

was found to be influenced by root activities (Objective 3), the mechanism of this effect is unclear.

In Objective 3, changes in dissolution and bioavailability in rhizosphere soil was hypothesized to be

a result of plant root exudates, and corresponding changes in rhizosphere soil pH. However, it could

also be a result of bacterial activity in the rhizosphere soil. Previous studies have shown that plant

root exudates affect bacterial activity and relevant enzyme activity22,23. Thus, a mechanistic

understanding of the role of root exudates or change in rhizosphere bacterial activity upon

- 88 -

dissolution of CuO NP in rhizosphere soil would require evaluating those influences separately.

Characterizing root exudation and bacterial community analysis (or enzyme activity) would be

required to answer this question. Such understanding would be helpful to more accurately model

and predict the dissolution behavior of metal or metal based ENMs in rhizosphere soil.

6.2.2. Optimize the way to measure toxicity of metal/metal oxide ENMs in soil. In Chapter 5, the dissolution profile of various Cu species correlated with their toxicity to

wheat plants. However, such correlation was based only upon four data points. To strengthen such

correlations, more experimental data are needed. Dissolution profiles for other Cu-based ENMs

(e.g. Cu(OH)2) and Cu ENMs containing product (e.g. Kocide® as Cu(OH)2, Nordox ® (Cu2O))

should be measured. Other types of biological endpoints (e.g. toxicity to other soil organisms, e.g.

soil isopods) should be measured.

Chapter 5 also suggested that the dissolution profile of different Cu-based ENMs should be

considered in the design of chronic toxicity tests for soil organisms. Usually, the length of a chronic

toxicity test is decided only by the tested organism. For example, the chronic toxicity test for Eisenia

andrei usually has 28 day period, while the chronic toxicity test for Hyalella Azteca usually has a 42 day

period24–26. There is currently no consideration of the NP properties on test design or duration. This

is a problem for those toxicity tests because the tested nanomaterials are transforming during the

test period (e.g. CuO and Cu(OH)2 NPs dissolve and release Cu2+) and the test could miss the salient

part of the exposure. Thus, future studies need to redefine the chronic toxicity testing protocols for

metal/metal oxide ENMs in soil, with the consideration of their dissolution profile. In this case, the

concept of ‘exposure profile’ explained in Chapter 5 could be used to better reflect the exposure of

the material instead of concentration.

6.2.3. Design ENMs that can solve the micronutrient deficiency problem in calcareous soil. One of the challenges for agriculture is micronutrient deficiency 27,28, especially in calcareous

(high pH) soils. One would predict that applying soluble ENMs to calcareous soil usually would not

solve the problem, due the low availability of metal ion at high pH27. In such cases, supplying

metal/metal oxide ENMs as foliar application and seed coatings might be better approaches to

preventing micronutrient deficiency. Foliar application with ENMs could be a feasible way of

efficiently delivering nutrients. Foliar applied ENMs have been shown to be able to enter and

translocate in plants29,30. Seed coatings could be devised to stick ENMs to the plant roots and slowly

release micronutrients to the rhizosphere during the plant growth period. However, there are still

- 89 -

challenges remaining: 1) To design ENMs that have the appropriate micronutrient release rate, in

which particle coatings, chemical composition, and size could be important tunable parameters. 2)

To monitor the bioavailability of relevant species in plant rhizosphere and the metal speciation

inside of plants. 3) To measure the availability of other metal/metal ENMs in the rhizosphere soil,

and to track the speciation of metal/metal ENMs inside plants.

6.3 References for Chapter 6 (1) Houba, V. J. G.; Novozamsky, I.; Lexmond, T. M.; Van der Lee, J. J. Applicability of 0.01 M

CaCl2 as a single extraction solution for the assessment of the nutrient status of soils and other diagnostic purposes. Commun. Soil Sci. Plant Anal. 1990, 21 (19–20), 2281–2290.




(5) Tourinho, P. S.; van Gestel, C. A.; Lofts, S.; Svendsen, C.; Soares, A. M.; Loureiro, S. Metal-based nanoparticles in soil: fate, behavior, and effects on soil invertebrates. Env. Toxicol Chem 2012, 31 (8), 1679–1692.

(6) Dimkpa, C. O.; Latta, D. E.; McLean, J. E.; Britt, D. W.; Boyanov, M. I.; Anderson, A. J. Fate of CuO and ZnO nano- and microparticles in the plant environment. Env. Sci Technol 2013, 47 (9), 4734–4742.

(7) Watson, J.-L.; Fang, T.; Dimkpa, C. O.; Britt, D. W.; McLean, J. E.; Jacobson, A.; Anderson, A. J. The phytotoxicity of ZnO nanoparticles on wheat varies with soil properties. Biometals 2015, 28 (1), 101–112.

(8) Kim, S. W.; Jung, J. H.; Lamsal, K.; Kim, Y. S.; Min, J. S.; Lee, Y. S. Antifungal effects of silver nanoparticles (AgNPs) against various plant pathogenic fungi. Mycobiology 2012, 40 (1), 53–58.

(9) Liu, R.; Lal, R. Potentials of engineered nanoparticles as fertilizers for increasing agronomic productions. Sci Total Env. 2015, 514, 131–139.

(10) Pradas del Real, A. E.; Castillo-Michel, H. A.; Kaegi, R.; Sinnet, B.; Magnin, V.; Findling, N.; Villanova, J.; Carriere, M.; Santaella, C.; Fernandez-Martinez, A. Fate of Ag-NPs in sewage sludge after application on agricultural soils. Env. Sci Technol 2016.


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(13) Peretyazhko, T. S.; Zhang, Q.; Colvin, V. L. Size-controlled dissolution of silver nanoparticles at neutral and acidic pH conditions: kinetics and size changes. Environ. Sci. Technol. 2014, 48 (20), 11954–11961.


(15) Baumann, E. W. Investigation of copper (II) chelates of EDTA and DTPA with cupric-selective electrodes. J. Inorg. Nucl. Chem. 1974, 36 (8), 1827–1832.

(16) Sillanpää, M.; Oikari, A. Assessing the impact of complexation by EDTA and DTPA on heavy metal toxicity using microtox bioassay. Chemosphere 1996, 32 (8), 1485–1497.

(17) Kapoor, S.; Mukherjee, T. Growth and reactivity of silver clusters in amino polycarboxylic acid solutions. J. Colloid Interface Sci. 2003, 264 (1), 301–306.

(18) Huang, S. S.; Liao, Q. L.; Hua, M.; Wu, X. M.; Bi, K. S.; Yan, C. Y.; Chen, B.; Zhang, X. Y. Survey of heavy metal pollution and assessment of agricultural soil in Yangzhong district, Jiangsu Province, China. Chemosphere 2007, 67 (11), 2148–2155.

(19) Fageria, N. K.; Slaton, N. A.; Baligar, V. C. Nutrient management for improving lowland rice productivity and sustainability. Adv. Agron. 2003, 80 (1), 63–152.

(20) Gambrell, R. P.; Wiesepape, J. B.; Patrick, W. H.; Duff, M. C. The effects of pH, redox, and salinity on metal release from a contaminated sediment. Water. Air. Soil Pollut. 1991, 57 (1), 359–367.

(21) Peng, C.; Xu, C.; Liu, Q.; Sun, L.; Luo, Y.; Shi, J. Fate and Transformation of CuO Nanoparticles in the Soil–Rice System during the Life Cycle of Rice Plants. Environ. Sci. Technol. 2017, 51 (9), 4907–4917.

(22) Baudoin, E.; Benizri, E.; Guckert, A. Impact of artificial root exudates on the bacterial community structure in bulk soil and maize rhizosphere. Soil Biol. Biochem. 2003, 35 (9), 1183–1192.

(23) Ai, C.; Liang, G.; Sun, J.; Wang, X.; Zhou, W. Responses of extracellular enzyme activities and microbial community in both the rhizosphere and bulk soil to long-term fertilization practices in a fluvo-aquic soil. Geoderma 2012, 173, 330–338.

(24) Ingersoll, C. G.; Brunson, E. L.; Dwyer, F. J.; Hardesty, D. K.; Kemble, N. E. Use of sublethal endpoints in sediment toxicity tests with the amphipod Hyalella azteca. Environ. Toxicol. Chem. An Int. J. 1998, 17 (8), 1508–1523.

(25) Ivey, C. D.; Ingersoll, C. G.; Brumbaugh, W. G.; Hammer, E. J.; Mount, D. R.; Hockett, J. R.; Norberg‐King, T. J.; Soucek, D.; Taylor, L. Using an interlaboratory study to revise methods for conducting 10‐d to 42‐d water or sediment toxicity tests with Hyalella azteca. Environ. Toxicol. Chem. 2016, 35 (10), 2439–2447.

(26) Robidoux, P. Y.; Svendsen, C.; Caumartin, J.; Hawari, J.; Ampleman, G.; Thiboutot, S.;

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Weeks, J. M.; Sunahara, G. I. Chronic toxicity of energetic compounds in soil determined using the earthworm (Eisenia andrei) reproduction test. Environ. Toxicol. Chem. An Int. J. 2000, 19 (7), 1764–1773.

(27) Chen, Y.; Barak, P. Iron nutrition of plants in calcareous soils. In Advances in agronomy; Elsevier, 1982; Vol. 35, pp 217–240.

(28) Rashid, A.; Ryan, J. Micronutrient constraints to crop production in soils with Mediterranean-type characteristics: a review. J. Plant Nutr. 2004, 27 (6), 959–975.

(29) Lowry, G. V; Avellan, A.; Gilbertson, L. M. Opportunities and challenges for nanotechnology in the agri-tech revolution. Nat. Nanotechnol. 2019, 14 (6), 517.

(30) Avellan, A.; Yun, J.; Zhang, Y.; Spielman-Sun, E.; Unrine, J. M.; Thieme, J.; Li, J.; Lombi, E.; Bland, G.; Lowry, G. V. Nanoparticle Size and Coating Chemistry Control Foliar Uptake Pathways, Translocation and Leaf-to-Rhizosphere Transport in Wheat. ACS Nano 2019.

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Appendices Appendix 1- Supporting information for Chapter 2: Develop a functional assay to measure the dissolution kinetics of metal-based nanoparticles in soil. Detailed nanoparticle amendment procedure

Nanoparticle powders were dispersed in 5mM NaHCO3 buffer in Milli-Q water (pH=7) (the final

particle concentration was 62 mg Cu/L for the low dose amendment and 620 mg Cu/L for the high

dose amendment). The suspension was then sonicated using an ultrasonic probe (Sonic

Dismembrators model 550, Fisher Scientific) for 30 s (We used a pulse mode with 5s pulse per 10 s

sonication, energy level 3, ~160W). The dispersed nanoparticles were then immediately added to

soil. Cu(NO3)2•2.5H2O powder was dissolved in Milli-Q water before being amended to the soil

(final concentration was 620 mg Cu/L for the high dose amendment, and 62 mg Cu/L for the low

dose amendment). For extraction experiments, for both Cu(NO3)2 and CuO NP amendments, 40ml

of the NP suspension or solution was added to 250g air-dried soil. For the control soil, we added 40

ml of bicarbonate buffer to 250 g of soil.

The amount of Milli-Q water added to each soil was controlled to ensure that each was at its field

capacity after amendment (16 wt%). For the high dose CuO NP amendment, Cu(NO3)2 amendment

and control amendment, after taking 10g of soil for digestion, the soil was further divided into 21

tubes (3 replicates * 7 different incubation periods). For the low dose CuO NP amendment and

Cu(NO3)2 amendment, after taking 10g of soil for digestion experiment, the soil was further divided

into 24 tubes (3 replicates * 8 different incubation periods).

For XAS study, for both Cu(NO3)2 and CuO NP amendments, 8ml of the NP suspension or

solution was added to 50 g air-dried soil (final moisture content was also 16%). For both high dose

and low dose amendment, after taking 10g of soil for the digestion experiment, soil was further

divided into 7 tubes (7 different incubation periods, no replicates).

All soil samples were incubated in 50ml centrifuge tubes under aerobic conditions for between 0 and

31 days before being extracted and digested. The moisture content of the soil samples was kept at

the soil field capacity by daily additions of Milli Q water.

- 93 -

Table A1-1. Calibration ranges used for ICP-MS measurement

Sample Calibration Range

DTPA extractable Cu in 10mg/kg

CuO NP amendment 0.01mg/kg-0.1mg/kg

CaCl2 extractable Cu in 10mg/kg

CuO NP amendment 0.002mg/kg – 0.01mg/kg


CuO NP amendment 0.05mg/kg-1mg/kg




Cu(NO3)2 amendment 0.01mg/kg-0.1mg/kg


Cu(NO3)2 amendment 0.01mg/kg-0.1mg/kg


Cu(NO3)2 amendment 0.05mg/kg-1mg/kg



Total Cu in 10mg/kg


Total Cu in 10mg/kg


Total Cu in 100mg/kg


Total Cu in 100mg/kg


All samples collected from

umamended soil 0.001mg/kg – 0.01mg/kg

- 94 -

Table A1-2. Total Cu measured in amended soils (4 replicates)

High dose CuO NP

amended soil (mg/kg)

(s.d.)

Low dose CuO NP


(s.d.)

High dose Cu(NO3)2


(s.d.)

Low dose Cu(NO3)2


(s.d.)

For extractions 99.6 (6.5) 13.1(0.7) 100.6 (9.4) 13.0 (0.6)

For XANES experiments

108.7 (1.3) 12.9 (0.1) 104.4 (1.7) 13.3 (0.2)

Table A1-3. Extractable Cu in unamended soil (3 replicates)

DTPA extraction CaCl2 extraction

Incubation time Wet soil extraction

(s.d.)

Dry soil extraction

(s.d.)

Wet soil extraction

(s.d.)

Dry soil extraction

(s.d.)

0 0.49 (0.04) 0.49 (0.04) 0.018 (0.010) 0.011 (0.002)

1 0.38 (0.17) 0.39 (0.08) 0.012 (0.001) 0.013 (0.002)

2 0.53 (0.12) 0.49 (0.03) 0.015 (0.006) 0.024 (0.012)

4 0.44 (0.03) 0.50 (0.00) 0.013 (0.009) 0.015 (0.007)

7 0.37 (0.05) 0.45 (0.01) 0.005 (0.002) 0.011 (0.003)

19 0.37 (0.01) 0.44 (0.01) 0.011 (0.002) 0.019 (0.006)

31 0.41 (0.01) 0.60 (0.14) 0.006 (0.001) 0.014 (0.004)

- 95 -

Figure A1-1. A) Primary particle size distribution determined from counting primary particles from 10 TEM imagines. B-K) Ten TEM images of CuO NP. Red bars indicate the counted particles. Heavily aggregated nanoparticles were not counted.

- 96 -

Figure A1-2. Size distribution of 80mg/kg CuO NP in pH=7, 5mM NaHCO3 buffer determined by

dynamic light scattering: (a) Number averaged size distribution, (b) intensity averaged size

distribution and (c) and autocorrelation function.

c

- 97 -

4 8 1 2

- 3 0

- 2 0

- 1 0

0

1 0

2 0

Ze

ta p

ote

nti

al

(mv

)

a )

p H

4 8 1 2

- 3 0

- 2 0

- 1 0

0

1 0

2 0

3 0

Ze

ta p

ote

nti

al

(mv

)

b )

p H

Figure A1-3. Zeta potential of 80mg/kg CuO NP as a function of pH measured in (a) 5mM

NaHCO3 buffer and (b) 5mM NaNO3. Error bars indicate ± 1standard error. The shift in the pH

of the isoelectric point from pH=8.8 in NaNO3 to pH=5.5 in bicarbonate indicates a specific

interaction between the CuO NPs and the carbonate species.

- 98 -

Figure A1-4. X-ray diffraction spectrum of CuO NP. The CuO NPs used here are identified as

tenorite.

12 22 32 42 52 62 72 82

Inte

ntis

ity

2 θ(degrees)

CuO NP Tenorite

- 99 -

Figure A1-5. pH of CaCl2 extracts in different amended and blank soils (pH values in CaCl2 extracts

of the high dose CuO NP amended soils at t=7 days after amendment were not measured.) Error

bars indicate ± 1standard error.

- 100 -

- 101 -

Figure A1-6. Extractable Cu and in wet and air dried amended soils as a function of time: (a)

DTPA extraction for 10 mg/kg amendment, (b) CaCl2 extraction for 10 mg/kg amendment, (c)

DTPA extraction for 100 mg/kg amendment and (d) CaCl2 extraction for 100 mg/kg amendment.

Error bars indicate ± 1 standard error. represents extractable Cu in CuO NP amended wet soils;

represents extractable Cu in CuO NP amended soils air dried after incubation; represents

extractable Cu in Cu(NO3)2 amended wet soils, and represents extractable Cu in Cu(NO3)2

amended soils air dried after incubation.

Extractions performed on wet and air-dried soils indicated that the influence of air-drying on DTPA

extractable Cu and CaCl2 extractable Cu in soils is insigficant for all amendments (P>0.05,

- 102 -

Kolmogorov–Smirnov test) (Figure S6). The insensitivity of CuO NP to air drying may result from

because the Cu(II) in the particles being oxygen-insensitive. The extractability of metal from oxygen-

sensitive (redox active) nanoparticles like metallic Cu NP or metallic Ag NP in soil may be more

affected by air drying than the CuO NP used here. This requires further study.

8 9 8 0 9 0 0 0 9 0 2 0 9 0 4 0 9 0 6 0

0

E n e r g y ( e v )

No

rma

liz

ed

ad

so

rpti

on

C u ( 0 )

C u S ( C o v e l l i t e )

C u ( I I ) N i t r a t e

C u ( I I ) P h o s p h a t e

C u ( I I ) S u l f a t e

C u O N P

C u ( I I ) - H A

C u ( I I ) F e r r i h y d r i t e

C u ( I ) - H A

C u 2 S ( C h a l c o c i t e )

Figure A1-7. XANES spectra for model compounds

Cu(II) sorbed to humic acid (Cu(II)-HA) was synthesized using established methods (Fulda et al.

2013)1. 118 mg of a humic acid standard (H1, 675-2, Sigma-Aldrich) was dissolved in 25 mL of DI

water, then 200 mg of CuCl2 was slowly added with stirring. The solution was filtered and dried.

- 103 -

The spectrum for Cu sorbed to ferrihydrite (Cu(II)- ferrihydrite) was obtained from Donner et al.

(2011)2. The Spectrum for Cu(I) sorbed to humic acid (Cu(I)-HA) was obtained from (Fulda et al.

2013)1.

Figure A1-7 indicated that reduction of Cu(II) to Cu(I) and sulfidation of Cu will result in a

significant white line shift, which is not observed in our experimental data. The LCF analysis

indicated CuO NP in soil transformed to Cu(II)-HA slowly over time.

Table A1-4. The results of the LCF analysis of the X-Ray Absorption Near Edge Structure

(XANES) region of the samples

Copper Time Goodness of fit

Percentages of Components

Sum R value Cu(II)- HA (%) CuO NP (%)

CuO NP Day 1 0.00022 14 86 100

CuO NP Day 4 0.00019 35 66 101

CuO NP Day 7 0.00014 41 59 100

CuO NP Day 19 0.00017 44 56 100

Cu(NO3)2 Day 0 0.00026 100 0 100

Cu(NO3)2 Day 1 0.00041 101 0 101

Cu(NO3)2 Day 4 0.00046 102 0 102

Cu(NO3)2 Day 7 0.00020 100 0 100

Cu(NO3)2 Day 19 0.00023 101 0 101

Appendix 2- Supporting information for chapter 3:Quantify the effect of soil properties, including soil moisture content, organic carbon content and pH, on the dissolution kinetics of CuO NP in soil.

Procedure for the extraction of SOM (FA and HA) from Lufa 2.1 soil

The SOM (here considered the fulvic acids (FA) and humic acids (HA) fractions) was extracted from

Lufa 2.1 soil following a procedure described by van Zomeren et al.3 A set of sixteen centrifuge tubes

containing the same amount of Lufa 2.1 soil were subjected to the same extraction procedure. Briefly,

40 mL of 0.1M HCl (pH 1.2) was added to 4 g of Lufa 2.1 soil (L/S=10) in a centrifuge tube, and the

suspension was mixed by continuous tumbling for 1h. Afterwards, the suspension was centrifuged

- 104 -

(3000g, 10 min). The supernatant was decanted and filtered using a 0.45 µm nitrocellulose membrane,

the pH adjusted to pH 2.0 with 1M NaOH and subsampled for dissolved organic carbon (DOC)

analysis. The remaining supernatant was stored frozen for further isolation from the inorganic matrix

(data not shown).

The soil residue was then neutralized with 0.1M NaOH, and thereafter 0.1M NaOH (pH 12) was

added under a N2 atmosphere to a final volume of 40 mL (L/S =10). The suspension was equilibrated

during 20h by intermittent shaking, and then centrifuged (3000g, 10 min), decanted and filtered, using

a 0.45 µm nitrocellulose membrane. The supernatant was acidified to pH 1.0 with 6M HCl and allowed

to stand overnight (≈20h) in a refrigerator (4ºC) to precipitate the HA fraction. The resultant

suspension (containing FA and inorganics) was centrifuged (3000g, 10 min), and the obtained

supernatant was decanted and filtered, the pH adjusted to 2.0 with 1M NaOH and stored frozen for

further extraction/isolation from the inorganic matrix. The precipitate with the HA fraction was then

washed with 10 mL of ultrapure water, the suspension was centrifuged (3000g, 10 min), and the

supernatant was discarded. The HA was then re-dissolved in a 0.1M NaOH solution, containing 0.2M

KCl, under a N2 atmosphere. The KCl was added to NaOH to increase the ionic strength, provoking

the flocculation of the colloidal inorganic material (non-humic material). The suspended solids were

removed by centrifugation (3000g, 10 min) and the HA solution was decanted under N2 atmosphere.

The process was repeated until practically all the HA was separated from the inorganic matrix. The

final solution of HA was acidified with 6M HCl to pH 1.0. The acidified solution was allowed to stand

overnight, and the supernatant was finally separated from the HA precipitate by centrifugation (3000g,

10 min). The HA was then re-dissolved in a minimum volume of 0.1M NaOH under a N2 atmosphere,

and then acidified to pH 2.0 and further desalted onto DAX-8 and cationic exchange resins.

The FA and HA fractions were isolated from the inorganic matrix by adsorption onto a Supelite™

DAX-8, followed by a cationic exchange resin in H+ mode, connected in series. The FA and HA

fractions, previously acidified to pH 2, were pumped through a DAX-8 column, and then the

inorganics were removed from the void volume with one column volume of ultra-pure water. The FA

and HA were then back eluted with 0.1M NaOH (3.5 column volumes) and directly transferred into

the cationic exchange column. Finally, the FA and HA solutions were freeze-dried and kept in a

dessicator over silica gel until further characterization by thermogravimetry and elemental analysis.

The DOC content of the influents and effluents from both resins were also measured for assessing

losses of organic carbon during the isolation/purification procedure.

- 105 -

Recovery of FA and HA from Lufa 2.1 soil and preliminary characterization

For a total LUFA 2.1 soil mass of 256 g, the total organic carbon extracted from the soil in each

fraction was: FA-SOM - 740.5 mg OC kg-1 soil d.w. and HA-SOM - 872.7 mg OC kg-1 soil d.w. The

total OC extracted from LUFA 2.1 soil was 1613.2 mg C kg-1, corresponding to an extraction efficiency

of 23.1 % of the total OC in the original soil sample. The FA-SOM and HA-SOM fractions account

for 46% and 54% of the total SOM extracted from LUFA 2.1 soil, respectively. The elemental analysis

data showed that the FA-SOM and HA-SOM fractions contain c.a. 39% and 32% carbon, respectively,

and that ash and impurities account for 24% and 38% of the sample mass, respectively.

Details on Cu speciation measurements

Cu speciation in Lufa soil and Arizona soil were analyzed by Cu K-edge x-ray absorption spectroscopy

at the Stanford synchrotron Radiation Lightsource (SSRL) on Beamlines 11-2 and 4-1, respectively.

Samples were ground, pressed into pellets, and placed between Kapton tape. Double crystal Si(220)

monochromator was calibrated by setting the first inflection of the K-edge of a metallic Cu foil to

8979 eV. On Beamline 11-2, harmonic rejection was achieved by use of a Rh-coated mirror, and

fluorescence data were recorded at 77 K using a 100-element germanium detector. On Beamline 4-1,

harmonic rejection was achieved by detuning the monochromator crystal by 20%, and fluorescence

data were recorded at 77 K using a 32-element germanium detector. The scans were averaged, energy

corrected using a metallic Cu foil standard and deadtime-corrected using SIXPack data analysis

software (v 1.4).4 Spectra background was subtracted and normalized before linear combination fitting

(LCF) analysis using Athena XAS data processing software (Demeter 0.9.24).5 A variety of spectra of

organic and inorganic Cu reference compounds were considered for LCF.6 Inclusion of a reference

spectrum into the combination fit required at least a 10% decrease in the Rf-value, indicating a

significant change to the quality of the fit.

Preliminary experiments

For all the CuO NP amendments using standard soils, preliminary experiments were conducted to

determine the approximate solubility of CuO NP in each soil. CuO NPs were added to soil to

achieve 50 mg/kg, 100 mg/kg and 500 mg/kg Cu. DTPA extractions were conducted after 30 days

of aging (moisture content was maintained at soil field capacity during aging period) to determine

the solubility of CuO NPs in each standard soil. The initial CuO NP concentration for each

- 106 -

treatment was then selected to ensure that the CuO NPs did not completely dissolve over the 30d

experiments.

Properties of soils (unamended) used in this study

Table A2-1: Properties of sampled soils

Soil Type pH Organic carbon

content (%)

Soil type Clay content

(%)

Background Cu

Lufa 2.1 soil 4.7 0.67 Silty sand 3.9 2.95mg/kg Lufa 2.2 soil 5.6 1.71 Loamy sand 8.3 3.4mg/kg Lufa 2.3 soil 5.9 0.66 Silty sand 7.6 3.3mg/kg Lufa 2.4 soil 7.4 1.99 Clayey loam 26.4 18.3mg/kg Arizona soil 7.6 0.54 Clayey loam 32.5 0.9 mg/kg Portugal soil 5.0 1.2 Sandy loam 3.1 9.1mg/kg

Table A2-2: Mass balance and experimental conditions for each treatment

Treatment

Targeted total Cu concentr

ation (mg/kg)

Total Cu measured from

digestion (mg/kg)

DTPA extractable Cu on

D30 (mg/kg)

CaCl2 extractable Cu on

D30 (mg/kg)

Residue Cu (non-extractable Cu2+) at D30 (mg/kg)

CuO NP remained at D30 (mg/kg)

pH Organic carbon content

(%)

Moisture content

Lufa 2.1 soil

100 114 75.0 4.9 18 21 5.0 0.67 16%

Lufa 2.1 soil, pH adjusted

100 112 1.30 0.050 0.33 110 7.4 0.67 16%

Lufa 2.1 soil, OM adjusted

300 287 190 12 47 49 4.9 1.34 16%

Lufa 2.2 Soil

500 503 249 2.6 62 192 5.9 1.71 21%

Lufa 2.2 Soil

250 265 171 1.3 43 54 5.8 1.71 21%

Lufa 2.2 Soil, pH adjusted

500 501 144 1.4 37 320 6.8 1.71 21%

Lufa 2.2 Soil,

moisture

500 481 234 2.3 58 189 5.9 1.71 10%

- 107 -

content adjusted Lufa 2.3

soil 500 539 68 0.37 17 455 6.5 0.66 17%

Lufa 2.4 soil

500 537 128 0.40 32 377 7.2 1.99 22%

Arizona soil

500 544 51 0.17 13 480 7.6 0.54 12%

Portugal soil

500 500 225 0.64 56 219 1.2 16% 16%

Nanoparticle characterization. The properties of the CuO NPs have been previously published6.

Briefly, the primary particle size was 38nm ± 1.7nm (TEM) and shape is spherical. The

hydrodynamic diameter of CuO NP was 560nm±103nm (pH=7), and the -potential was -

16.1mV±1.7mV. The isoelectric point of CuO NP is 8.8 (in 5mM NaNO3), so all CuO NPs are

positively charged in soil pore water in all the treatments.

Derivation of dissolution model

The transformation of the CuO NPs in the soil is given by equation A2-1:

CuO NP(s) ⇄ 𝐶𝐶𝐶𝐶2+𝑑𝑑𝐸𝐸𝐶𝐶𝑓𝑓𝑓𝑓𝑓𝑓𝐸𝐸𝑟𝑟𝑓𝑓𝐶𝐶𝐸𝐸,𝐾𝐾𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙��𝐶𝐶𝐶𝐶[𝐿𝐿] (𝐴𝐴2 − 1)

Where L represents various ligands that can complex with Cu.

According to the assumptions in our manuscript:


= 𝑘𝑘𝑑𝑑,𝐷𝐷𝐴𝐴𝐶𝐶𝐶𝐶𝐶𝐶 − 𝑘𝑘𝑟𝑟,𝐷𝐷[𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑜𝑜𝑑𝑑,𝑑𝑑1

1+𝐾𝐾𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝐴𝐴𝐶𝐶𝐶𝐶𝐶𝐶 (𝐴𝐴2 − 2)

Where the subscript 𝐴𝐴𝐶𝐶𝐶𝐶𝐶𝐶 indicates the total surface area (not mass) of CuO NP. 𝑘𝑘𝑑𝑑,𝐷𝐷 ,𝑘𝑘𝑟𝑟,𝐷𝐷 are the

dissolution rate constant and reverse reaction (precipitation) constant with respect to NP surface

area. 𝐾𝐾𝑓𝑓𝑓𝑓𝑙𝑙𝑓𝑓𝑓𝑓𝑑𝑑 is the partitioning constant between Cu associated with natural ligands (including both

DOM and soil surfaces, e.g. SOM, clay, iron oxides) and free Cu2+.

𝐴𝐴𝐶𝐶𝐶𝐶𝐶𝐶 = 𝐶𝐶1 ∗ �[𝐶𝐶𝐶𝐶𝐶𝐶]𝑇𝑇𝜌𝜌𝐶𝐶𝐶𝐶𝐶𝐶

�𝑓𝑓

= 𝐶𝐶2 ∗ [𝐶𝐶𝐶𝐶𝑂𝑂]𝑑𝑑2/3 (𝐴𝐴2 − 3)

- 108 -

𝜌𝜌𝐶𝐶𝐶𝐶𝑂𝑂 is the density of CuO NP. For spherical particles (e.g. the CuO NP used in this study),

n=2/3.7 The constant C1 takes the initial size of the particles into account. Assuming density and

shape of the particle would not change during the dissolution process, then C1 and 𝜌𝜌𝐶𝐶𝐶𝐶𝑂𝑂 can be

incorporated into C2, which should be a constant during the dissolution process. Thus, equation S3-

2 can be rewritten as:


= 𝑘𝑘𝑑𝑑[𝐶𝐶𝐶𝐶𝑂𝑂]𝑑𝑑2/3 − 𝑘𝑘𝑟𝑟[𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑜𝑜𝑑𝑑,𝑑𝑑

11+𝐾𝐾𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙

[𝐶𝐶𝐶𝐶𝑂𝑂]𝑑𝑑2/3 (𝐴𝐴2 − 4)

Where the exponent 2/3 converts concentration to surface area; 𝑘𝑘𝑑𝑑 𝑎𝑎nd 𝑘𝑘 𝑟𝑟 are the rate constants

and are fitted parameters in this model. 𝐾𝐾𝑓𝑓𝑓𝑓𝑙𝑙𝑓𝑓𝑓𝑓𝑑𝑑 is determined from the CuSO4 control experiment

as described in the “Dissolution models” section in the manuscript. Note, [𝐶𝐶𝐶𝐶𝑂𝑂]𝑑𝑑 = [𝐶𝐶𝐶𝐶𝑂𝑂]𝑜𝑜 −

[𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑜𝑜𝑑𝑑,𝑑𝑑

The following equation was used to numerically fit the experimental data:

[𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑜𝑜𝑑𝑑,𝑑𝑑+1= [𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑜𝑜𝑑𝑑,𝑑𝑑+t*( 𝑘𝑘𝑑𝑑[𝐶𝐶𝐶𝐶𝑂𝑂]𝑑𝑑2/3 − [𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑜𝑜𝑑𝑑,𝑑𝑑

𝑘𝑘𝑟𝑟1+𝐾𝐾𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙

[𝐶𝐶𝐶𝐶𝑂𝑂]𝑑𝑑2/3) (A2-5)

Equation A2-5 was used to fit the experimental data to get kd and kr.

- 109 -

Table A2-3. Comparison between CuO NP dissolution measured by XANES and chemical

extraction

The details on XANES fitting result can be found in table A2-4 and Figure A2-1. XANES could not

detect the minimal (<10%) CuO NP dissolution in the high pH Arizona soil. In Arizona soil, with

500mg/kg CuSO4 dosing, ~20% of Cu precipitated as copper carbonate, with most Cu (~80%)

remaining in its complexed form (not precipitating as a new solid phase). This suggested that in the

CuO NP treatment, with only 64mg/Kg Cu dissolved, the amount of Cu ions released from CuO NP

would not precipitate as other Cu phases as the concentration of Cu ions released from CuO NP did

not exceed the solubility limit of Cu with respect to Copper carbonate in Arizona soil .

Treatment % CuO NP dissolved by

XANES (mg/kg)

% CuO NP dissolved by

extraction(mg/kg)

Lufa 2.2 soil, 500 mg/kg

treatment, 10% moisture,

D7

37 32

Lufa 2.2 soil, 500mg/kg

treatment,10% moisture

D31

52 61

Lufa 2.2 soil, 500mg/kg

treatment, 21% moisture

D31

60 62

Arizona soil, 300mg/kg ,

12% moisture content D0

0 1

Arizona soil, 300mg/kg ,

12% moisture content D7

0 9

- 110 -

Table A2-4. Linear combination fitting results of k3-weighted Cu EXAFS spectra (Figure S3-1) for

Arizona soil exposed to 300mg/kg of CuO NP or CuSO4. Samples were fit over a k range of 3-12 Å.

The percentages have inherent ±15% uncertainties. Data are presented with the R factor (Rf) and the

Reduced χ2 parameters to indicate the quality of the fits.

Reference Compounds Fit Parameters

CuO NP

(%)

Cu-

NOM

(%)

Cu-Ferri

(%)

Cu-Carb

(%) R factor Red. χ2

CuO

NP

Day 0 119 - - - 0.03 0.65

Day 7 109 - - - 0.03 0.56

Day 21 110 - - - 0.03 0.52

CuSO4 Day 0 - 44 41 19 0.08 0.60

Day 7 - 60 23 20 0.10 0.67

- 111 -

Figure A2-1. Cu EXAFS spectra (black) and linear combination fits (red) for CuO NP and CuSO4

exposed soil. (a) Arizona soil, (b) Lufa 2.2 soil. Cu-Ferrihydrite and Cu-NOM models were

synthesized using established methods.1,8

- 112 -

0 1 0 2 0 3 0 4 0

0

2 0

4 0

6 0

8 0

1 0 0

T i m e ( d a y s )E

xtr

ac

tab

le C

u

(mg

/k

g d

rie

d s

oil

)

D T P A e x t r a c t i o n , L u f a 2 . 1 s o i l

p H 5 . 0

p H 7 . 4

Figure A2-2. DTPA extractable Cu in Lufa 2.1 soils dosed with 100mg/kg CuO NPs at pH 5.0

(squares) and pH 7.4 (triangles). Bars are standard deviation of the measurements.

Figure A2-2 shows that the overall dissolution of CuO NPs in Lufa 2.1 soil was greatly reduced by

increasing soil pH from 5.0 to 7.4. However, due to the very slow dissolution, the dissolution

kinetics of CuO NPs in pH 7.4 soil could not be accurately modeled, thus the data was not used to

calculate an dissolution rate constant and solubility.

- 113 -

Cross validation of the pH- kd correlation

0 . 0 0 0 0 0 0 0 . 0 0 0 0 0 5 0 . 0 0 0 0 1 0 0 . 0 0 0 0 1 5

0 . 0

0 . 5

1 . 0

1 . 5

T a k e o u t L u f a 2 . 1

{ H+

} ( m o l / L )

kd

(mg

1/3

·kg

1/3

·s-1

)

0 1 0 2 0 3 0 4 0

0

5 0

1 0 0

1 5 0

T i m e ( d a y s )

Dis

so

lve

d C

u (

mg

/kg

)


P r e d i t i o n u p p e r 9 5 % C I


0 . 0 0 0 0 0 0 0 . 0 0 0 0 0 5 0 . 0 0 0 0 1 0 0 . 0 0 0 0 1 5

0 . 0

0 . 5

1 . 0

1 . 5

T a k e o u t L u f a 2 . 2 2 1 %

{ H+

} ( m o l / L )

kd

(mg

1/3

·kg

1/3

·s-1

)

0 1 0 2 0 3 0 4 0

0

1 0 0

2 0 0

3 0 0

4 0 0

T i m e ( d a y s )

Dis

so

lve

d C

u (

mg

/kg

)




0 . 0 0 0 0 0 0 0 . 0 0 0 0 0 5 0 . 0 0 0 0 1 0 0 . 0 0 0 0 1 5

0 . 0

0 . 5

1 . 0

1 . 5

T a k e o u t l u f a 2 . 2 2 5 0 p p m

{ H+

} ( m o l / L )

kd

(mg

1/3

·kg

1/3

·s-1

)

0 1 0 2 0 3 0 4 0

0

1 0 0

2 0 0

3 0 0

T i m e ( d a y s )

Dis

so

lve

d C

u (

mg

/kg

)




0 . 0 0 0 0 0 0 0 . 0 0 0 0 0 5 0 . 0 0 0 0 1 0 0 . 0 0 0 0 1 5

0 . 0

0 . 5

1 . 0

1 . 5

T a k e o u t L u f a 2 . 2 1 0 %

{ H+

} ( m o l / L )

kd

(mg

1/3

·kg

1/3

·s-1

)

0 1 0 2 0 3 0 4 0

0

1 0 0

2 0 0

3 0 0

4 0 0

T i m e ( d a y s )

Dis

so

lve

d C

u (

mg

/kg

)




0 . 0 0 0 0 0 0 0 . 0 0 0 0 0 5 0 . 0 0 0 0 1 0 0 . 0 0 0 0 1 5 0 . 0 0 0 0 2 0

0 . 0

0 . 5

1 . 0

1 . 5

2 . 0

T a k e o u t L u f a 2 . 1 w i t h O M a j u d s t e d

{ H+

} ( m o l / L )

kd

(mg

1/3

·kg

1/3

·s-1

)

0 1 0 2 0 3 0 4 0

0

1 0 0

2 0 0

3 0 0

T i m e ( d a y s )

Dis

so

lve

d C

u (

mg

/kg

)




Figure A2-3: Cross validation of the correlation between kd and {H+}. The figures on the left are

the correlations after taking out one soil treatment, the figures on the right are the resulting

predictions on the removed soils.

- 114 -

Multivariate regressions:

Table A2-5.Multivariate regression between dissolution rate constant and soil organic matter content and hydrogen ion activity.

Coefficients Standard

Error P-value Lower 95%

Upper 95%

Organic carbon content 0.119 0.185 0.55 -0.356 0.594 {H+} 5.503*105 1.733*105 0.025 1.05*105 9.96*105

Table A2-6.Multivariate regression between solubility and soil organic matter content and hydrogen ion activity.


Error P-value Lower 95% Upper 95%

Organic carbon content 210.6 35.1 0.0018 120 301 {H+} 8.40*106 3.29*106 0.81 -7.62*106 9.30*106

Table A2-7.Multivariate regression between reverse reaction rate constant and soil organic matter content and hydrogen ion activity.


Error P-value Lower 95% Upper 95%

Organic carbon content -0.108 0.0600 0.13 -0.261 0.0456 {H+} -9.17*103 5.60*103 0.16 -2.36*103 5.25*103

As suggested from Table A2-5 and Table A2-6, solubility is correlated with only organic matter

content (P<0.05), and that dissolution rate is correlated with only {H+}(P<0.05). Table A2-7

indicates that kr shows no correlation with either organic matter content or {H+}.

- 115 -

Appendix 3- Supporting information for Chapter 4: Elucidate CuO NP dissolution behavior and toxicity to wheat (Triticum aestivum) in rhizosphere and bulk soil.

Table A3-1: Total Cu concentration (mean (SD), mg/kg) in soil for each treatmentga

CuO NP CuSO4 Control

Fresh 534.9 (39.4) 307.1 (3.7) 5.9 (0.3)

Aged 510.5 (6.4) 312.5 (1.3) N.A.

Table A3-2: DTPA extractable Cu (mean (SD), mg/kg) in the control treatment before and after plant growth

Bulk soil Loosely attached

soil

Rhizosphere soil

day 0 1.3 (0.1) N.A. N.A.

day 14 1.5 (0.1) 1.6 (0.4) 1.3 (0.9)

- 116 -

Figure A3-1. Change in DTPA extractable Cu and CaCl2 extractable Cu for 250mg/kg CuO NP

treatment, 250mg/kg and 500mg/kg CuSO4 treatments (without growing plants) over 30 days. Error

bars are standard deviations.

A decreasing trend in both CaCl2 extractable Cu and DTPA extractable Cu were observed in CuSO4

treatments (both 500mg/kg and 250 mg/kg), while an increasing trend in both CaCl2 extractable Cu

and DTPA extractable Cu were observed in 500mg/kg CuO NP treatment. DTPA can extract

~80% of Cu for both 500mg/kg and 250mg/kg CuSO4. The solubility of CuO NP was estimated to

be 312mg/kg by fitting a first-order dissolution kinetic model to the data (R2=0.990) and assuming

80% of the Cu was extractable.

- 117 -

Figure A3-2. Different soil regions defined in this study9,10. Bulk soil was presumed to be unaffected

by the roots or root exudates due to lack of proximity to the roots. Loosely attached soil clung to

roots but could be shaken free. Rhizosphere soil was shaken off of the roots after air drying.

Nanoparticle characterization6. Primary particle size distribution of the CuO NP was previously

measured by transmission electron microscopy (TEM, Hitachi H-9000 TEM microscope operating at

300 kV). Particles in 10 TEM imagines were counted for primary particle size distribution 278

nanoparticles were counted). The hydrodynamic diameter and zeta potential of CuO NP (80 mg/kg

as Cu) in 5mM NaHCO3 buffer (pH=7) were measured by dynamic light scattering (Zetasizer Nano,

Malvern) 6. The isoelectric points of CuO NP (80mg/kg) in 5mM NaHCO3 buffer and in 5mM

NaNO3 were calculated from measurements of the zeta potential of the particles in suspension over

a range of pH6.

Soil amendment. A nanoparticle suspension containing 644mg CuO NP, 664mg Na2SO4, and

208ml milli-q water (CuO NP treatment), dissolved Cu2+ solution containing 746mg CuSO4 and

208ml milli-q water (CuSO4 treatment), or Na2SO4 solution (664mg Na2SO4 and 208ml milli-q water

(control treatment) was sonicated (Sonic Dismembrators model 550, Energy level 3) for 30s before

adding to 1000kg air dried Lufa 2.2 soil. The Na2SO4 treatment was used to control for any effect of

added SO4 ions to plants.

- 118 -

The soil-suspension mixture (for NPs) or soil-solution mixture (for dissolved species) was

hand-mixed using a polycarbonate spoon for 20 min for each treatment. The soil was aged in an

incubator, and the moisture content was maintained at field capacity by daily addition of milli-Q-

water. For the control (no Cu added), only the fresh treatment protocol was followed.

Soil and Plant digestion. The soil samples collected before the plant growth experiment were

analyzed for total Cu content to confirm the amount of Cu in each treatment11. Briefly, 1g of air-dried

soil was digested with concentrated nitric acid and 30% hydrogen peroxide at 95 ºC using a hot plate.

After digestion, the samples were centrifuged at 3000 rpm for 10 min, followed by filtration using a

0.45um filter to remove fine particles in the supernatant. The filtered supernatant was diluted with

Milli-Q water and acidified with 20% HNO3 (final HNO3 concentration was 2%) for ICP-MS (Agilent

7700x) analysis. The plant total Cu concentration was determined using acid digestion according to a

modified version of U.S. EPA Method 3050B (1996).11 The dried root and shoot tissues were digested

with concentrated nitric acid and 30% hydrogen peroxide at 95°C using a hot plate. Due to the small

sample size, the chemical additions were 1/10th the volumes listed in EPA Method 3050B11. After

digestion, the samples were centrifuged at 3000 rpm for 15 min and the supernatant was diluted with

Milli-Q water to a final HNO3 concentration of ~5% for analysis by ICP- MS (Agilent 7700x).

ICP-MS measurement. Germanium was used as an internal standard for quality control. The range

of the calibration curve used in each treatment was between 0.005mg Cu/kg to 0.5mgCu/kg. All the

extraction samples, digestion samples were diluted within the range of the calibration curve. All the

samples were diluted and acidified right after sample collection, and measured within two days after

preparation. The calibration samples were measured each time when measuring extraction/digestion

samples (with all the calibration curves measured, R2 values were > 0.998 at all times).

Cytoviva analysis. The interaction between roots and NPs were visualized in fresh roots after a

rinsing step in 10-3 M KCl, using a DF-HSI system (CytoViva Inc., USA). This enhanced resolution

dark–field microscope system (BX51, Olympus, USA) was equipped with a 150 W halogen light source

for the dark–field sample illumination (Fiber-Lite®, Dolan-Jenner, USA), and a hyperspectral camera

(CytoViva Hyperspectral Imaging System 1.4). The roots were observed with 600× and 1000×

magnification. Hyperspectral images were acquired using 60% light source intensity and 0.1 s

acquisition time per line. Each pixel of the hyperspectral image contains its light reflectance spectrum

ranging from 400 to 1000 nm with a step of 1.5 nm.

- 119 -

Hyperspectral images (datacubes) were acquired on samples of CuO NPs mixed with Lufa 2.1

soil for 15min using ENVI 5.2 software (Exelis Visual Information Solutions, CO, United States).

After normalizing the datacubes for the lamp signal, spectral pre-libraries of CuO NPs in gels were

collected from these reconstructed RGB images based on pixel purity index as described in an earlier

article12. The spectra contained in the pre-library that were not specific to the CuO NP were filtered

by matching the datacubes against negative controls (soil without CuO NP, and control roots) using

a Spectral Angular Mapping algorithm (SAM, ENVI 5.2), an algorithm comparing angles between

vectors. Spectra in the pre-library that matched spectra of pixels in the control hyperspectral images

were considered as unspecific false positives and removed from the pre-library, whereby two vectors

(i.e. spectra) with angles ≤0.085 rad were considered as similar. The remaining spectra constitute the

final CuO NP library (i.e. exclusively containing specific hyperspectral CuO NP signature). The spectral

library of soil only (background), CuO NP spectral libraries, and the SAM results to test their

specificity on negative controls are shown in Figure S3.

The CuO NP library was used to perform SAM on hyperspectral images of exposed and

control roots, using the threshold angle of 0.085 rad, with two replicate pictures per condition on

several root areas. Each pixel in the images matching the hyperspectral signature of CuO NPs was

highlighted in red.

- 120 -

Figure A3-3. (A) Spectral library of the CuO-NPs. The spectral library has been built using

datacubes of CuO mixed with hydrated soil. (B) Example of SAM (Spectral Angle Mapping) results

to test for the specificity of the spectral library using positive controls (soil containing CuO NPs) or

negative controls (soil without CuO NPs or control root) images. The pixels containing the spectral

signal of CuO NP are highlighted in red (bottom line). Note that only the positive control contained

the signal of CuO-NP.

- 121 -

Figure A3-4. DTPA extractable Cu on bulk soil and bulk bottom soil in different Cu treatments

In all treatments, no significant differences (P<0.05, unpaired t-test) were found between DTPA

extractable Cu in bulk soil and bulk bottom soil, suggesting no vertical transport of Cu in all

treatments. Error bars are one standard deviation.

0

100

200

300

CuO Fresh CuO Aged CuSO4 Fresh CuSO4 Aged

Ext

ract

able

Cu

conc

entr

atio

n (m

g/kg

)DTPA extraction

bulk soil

bulkbottom soil

- 122 -

Table A3-3: Samples that provided sufficient soil for DTPA extraction for rhizosphere soil and loosely attached soil

Rhizosphere soil Loosely

attached soil

Plant number

CuO NP Fresh

CuO NP aged

Control CuO NP Fresh

CuO NP aged

Control

1

X X

X X 2 X X X X 3 X

X X

X

4

X X X 5 X

X X X X

6

X X X X

Table A3-4: Samples that provided sufficient soils for CaCl2 extraction for rhizosphere soil and loosely attached soil

Rhizosphere soil Loosely attached soil

Plant number

CuO NP Fresh

CuO NP aged

Control CuO NP Fresh

CuO NP aged

Control

1 X

X

X X 2

X X X X

3 X

X X

X 4 X

X X

5

X X X X 6

X X X X X

- 123 -

Figure A3-5. Representative photos showing Cu toxicity led to shortened root and/or root

compactness in fresh CuSO4 treatment (a) aged CuSO4 treatment (b) and aged CuO NP treatment

(d), whereas these toxic effects were not observed in fresh CuO NP treatments (c) and in control

treatments (e).

a b

c d e

- 124 -

Figure A3-6. Hyperspectral imaging of plant roots grown in soil with CuO-NP, CuSO4 or Na2SO4

(control) freshly amended or after aging. Roots exposed to CuSO4 (both after soil aging or not)

showed a brown-damaged (necrotic) zone, that was not found on any of the CuO NP exposed

roots.

- 125 -

Figure A3-7. Mean concentration of Cu (mg/kg) in wheat tissue (dry weight): a) Cu concentration

in shoots, b) Cu concentration in roots. Error bars show ± 1 SD. Capital letters indicate significant

different differences between groups (one way ANOVA test, P≤0.05).

Appendix 4. Explanation on the solubility of CuO NP in Chapter 3

When reviewing Chapter 3 of this thesis, we recognized an inconsistency in the data set: In section

3.2.6, 500mg/kg CuSO4 was shown to be completely soluble in Lufa 2.2 soil, i.e. having no

observable Cu precipitation (<~5 wt%) as suggested from the XANES spectra. However, the

modeled solubility of CuO NP in Lufa 2.2 soil was ~300mg/kg. For the equilibrium condition to

exist, these two scenarios (adding the same mass of either CuO NPs of CuSO4 salt) should have led

to the same final condition, but they did not. One reasonable explanation is presented schematically

in Figure A4-1. The CuSO4 treatment involved missing of Cu ions thoroughly through the soil,

providing a high degree of mixing between Cu and the soil NOM (Figure A4-1, b). In contrast, the

CuO NPs are not mixed as homogeneously in soil compared to CuSO4 treatment (Figure A4-1,a).

Instead, there is a localized dissolution-sorption equilibrium near the CuO NPs, and therefore not all

of the soil NOM is contacted with Cu. Thus, in CuO NP treatment, there is a fraction of SOM that

is not occupied by the Cu ions released by CuO NPs when the localized solubility of CuO NP has

been reached, resulting a less solubility of CuO NP compared to the sorption capacity of Cu in soil.

- 126 -

Figure A4-1. Conceptual model of Cu speciation in Lufa 2.2 soil a) CuO NP treatment b) CuSO4

treatment. The black dots in a) represents CuO NPs, and the shadowed brown area suggest Cu-OM

complex. The Cu-OM complex on b) is evenly distributed in soil, whereas in a) it is only evenly

distributed around particles. Figure A4-1. Conceptual model of Cu speciation in Lufa 2.2 soil a)

CuO NP treatment b) CuSO4 treatment. The black dots in a) represent CuO NPs, and the

shadowed brown area represents the localize area around the CuO NPs where the released Cu has

formed Cu-OM complexes. The Cu-OM complex in b) is evenly distributed in soil because the

added Cu ion can be uniformly mixed into the soil, contacting all of the soil NOM. While the CuO

NPs are uniformly distributed in soil, the released ions do to contact all of the soil NOM. This

lowers the overall "apparent" solubility in soil relative to the case for Cu ion addition.

Reference for Appendices: (1) Fulda, B.; Voegelin, A.; Ehlert, K.; Kretzschmar, R. Redox transformation, solid phase speciation and solution dynamics of copper during soil reduction and reoxidation as affected by sulfate availability. Geochim. Cosmochim. Acta 2013, 123, 385–402.

(2) Donner, E.; Howard, D. L.; Jonge, M. D. de; Paterson, D.; Cheah, M. H.; Naidu, R.; Lombi, E. X-ray absorption and micro X-ray fluorescence spectroscopy investigation of copper and zinc speciation in biosolids. Env. Sci Technol 2011, 45 (17), 7249–7257.


(4) Webb, S. M. SIXpack: a graphical user interface for XAS analysis using IFEFFIT. Phys. Scr. 2005, 2005 (T115), 1011.

(5) Ravel, B.; Newville, M. ATHENA, ARTEMIS, HEPHAESTUS: Data analysis for X-ray absorption spectroscopy using IFEFFIT. J. Synchrotron Radiat. 2005, 12 (4), 537–541.

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