Formation of Particulate Matter from the Oxidation of Evaporated

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Energy and the Environment

Formation of Particulate Matter from the Oxidation of Evaporated Hydraulic Fracturing Wastewater Jeffrey Kevin Bean, Sahil Bhandari, Anthony Bilotto, and Lea Hildebrandt Ruiz Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.7b06009 • Publication Date (Web): 29 Mar 2018 Downloaded from http://pubs.acs.org on March 30, 2018

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Environmental Science & Technology

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Formation of Particulate Matter from the Oxidation of Evaporated Hydraulic

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Fracturing Wastewater

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Jeffrey K. Bean1, Sahil Bhandari1, Anthony Bilotto2, and L. Hildebrandt Ruiz1*

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1

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Texas, 78712

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* Correspondence to L. Hildebrandt Ruiz ([email protected])

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Abstract

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The use of hydraulic fracturing for production of petroleum and natural gas has increased

McKetta Department of Chemical Engineering, The University of Texas at Austin, Austin,

Covenant Testing Technologies LLC, Sugar Land, Texas, 77478

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dramatically in the last decade, but the environmental impacts of this technology remain unclear.

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Experiments were conducted to quantify airborne emissions from twelve samples of hydraulic

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fracturing flowback wastewater collected in the Permian Basin, as well as the photochemical

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processing of these emissions leading to the formation of particulate matter. The concentration of

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total volatile carbon (TVC, hydrocarbons evaporating at room temperature) averaged 29 mg of

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carbon (C) L-1. After photochemical oxidation under high NOx conditions the amount of organic

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particulate matter (PM) formed per milliliter of wastewater evaporated averaged 24 µg; the

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amount of ammonium nitrate formed averaged 262 µg. Based on the mean PM formation

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observed in these experiments, the estimated formation of PM from evaporated flowback

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wastewater in the state of Texas is in the range of estimated PM emissions from diesel engines

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used in oil rigs. Evaporation of flowback wastewater, a hitherto unrecognized source of

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secondary pollutants, could significantly contribute to ambient PM concentrations.

ACS Paragon Plus Environment


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1

Introduction

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The production of natural gas and petroleum in the United States has increased dramatically in

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recent years due to new technologies including horizontal drilling and hydraulic fracturing which

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greatly decrease the cost to harvest gas and oil from shale formations. In 2011 the United States

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first became the world’s largest producer of natural gas; during that year, gas from shale

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formations accounted for 34% of the total U.S. production.1 The environmental impacts of this

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unconventional natural gas development (UNGD), including its impacts on air quality, are

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largely unknown. Several studies have estimated airborne emissions from hydraulic fracturing

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activity based on bottom-up calculations 2 and ambient measurements 3–5 including

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comprehensive measurements focusing on methane. 4,5 However, a full classification of airborne

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emissions from UNGD does not exist.6

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A key difference between hydraulic fracturing and traditional oil and gas production is the use of

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high volumes of fluid to fracture geologic formations. The proprietary composition of the

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fracturing fluid varies between drilling companies and can also depend on the type and depth of

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reservoir formation. The bulk of the fluid is water (and suspended sand), composing up to 99%

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of the fluid, but it typically also includes a long list of additives.7 As of 2013, the US EPA had

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catalogued 42 purposes of additive chemicals, including acids for well cleaning, proppants for

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opening and growing fractures, and biocides for preventing and controlling microorganism

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growth. 7,8 Several studies have been published on hydraulic fracturing chemicals, their purpose

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and their toxicity based on public and commercial databases including the US hydraulic

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fracturing chemical registry, FracFocus. 7–12 These studies have also shown a lack of available

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physicochemical data for fracturing chemicals; for example, an EPA database of 1173 chemicals

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lacks measured volatility data (Henry’s Law Constants) for over 80% of the chemicals.13 Among


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identified organic compounds, alcohols, petroleum hydrocarbons, and small aromatic compounds

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are the dominant classes and are of concern as primary pollutants 14,15 as well as precursors to

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secondary pollutants. As for total organic carbon content, Hayes 16 reported that freshly prepared

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fracturing fluid in the Marcellus contained as much as 1300 mg L-1 dissolved organic carbon

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with a median 226 mg L-1 (from 19 locations).

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The hydraulic fracturing of a horizontally drilled well in a shale formation can require 10 million

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gallons of water 17,18 in the Permian basin an average of 8 million gallons of water are used per

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well. During the first year of production approximately 50% of this water turns to the surface as

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flowback fluid. Over the life of a well the volume of flowback fluid in the Permian Basin often

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exceeds the volume of original fracturing fluid. 18

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Flowback wastewater is a mix of the original fracturing fluid and produced water - the water

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from oil and gas wells which originates in geologic formations and constitutes the largest waste

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stream generated in oil and gas production. 19 Management of flowback fluid has traditionally

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varied by state, basin, and at times, operator. EPA regulations (NSPS OOOOa20) mandate that

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operators maximize their use of a reduced emissions completions layout, which varies based on

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two stages of flowback: initial stage and separation stage. During the initial stage, flowback fluid

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can route to an open-air fracturing tank, lined pit, or other vessel. Most operators first route the

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fluid through a sand knockout to reduce the eroding effects of high velocity sand on equipment,

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followed by a pressure control device (manifold), and then the temporary storage vessel. Once

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enough gas is present, the fluid is sent to a 3 or 4-phase separator and divided into sand,

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condensate (oil), gas (lighter hydrocarbons), and wastewater. After separation, each portion of

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the fluid is routed to a respective storage tank for sale or disposal. In the United States,

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condensate and wastewater are usually stored in temporary tanks8, though flowback wastewater



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can also be directed to holding ponds which freely evaporate to the atmosphere. According to a

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recent national estimate based on state and federal data, about 3.6% of all produced water in

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United States in the year 2012 was managed via evaporation, totaling about 29 billion gallons of

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water. 19

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Volatile compounds in the wastewater can evaporate and affect atmospheric composition. These

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potential emissions and their effects on air quality have not received much attention and are the

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focus of this work. Organic compounds which evaporate into the atmosphere can be oxidized

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and contribute to the formation of ozone and particulate matter (PM). Ozone formation is mostly

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influenced by lighter compounds which form a significant fraction of emissions but, due to their

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high vapor pressure, do not persist in flowback wastewater. Larger compounds persist in the

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flowback wastewater for longer periods and have higher potential to form particulate matter.

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While several studies report PM levels near and on-site oil and natural gas processes, attribution

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of PM formation to specific processes and/or primary chemicals has been limited.6

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In this work, twelve samples of flowback wastewater collected from separators (n = 4) and

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flowback storage tanks (n = 8) in the Permian Basin were analyzed for their total volatile carbon

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(TVC) content and TVC emission rates. Laboratory chamber experiments were conducted to

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oxidize wastewater emissions and quantify the resulting formation of particulate matter.

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Wastewater evaporation and subsequent atmospheric processing is a hitherto unrecognized

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source of PM.

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2

Materials and Methods

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Twelve samples of hydraulic fracturing flowback wastewater were collected from the Wolfcamp

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shale formation in the Permian Basin. All samples were collected during the flowback process

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from wells that produced both oil and gas, either directly after the 4-phase separator (samples 1,

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4, 7 and 10) or from an open-top flowback tank (all other samples). Additional details on sample

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origin, collection and storage are provided in section S1 of the supplemental information and are

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summarized in Table 1.

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2.1

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Total volatile carbon (TVC, operationally defined) of each sample was measured by allowing

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0.2-0.8 mL of wastewater sample to evaporate completely in 40-60 L of clean air from an Aadco

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clean air generator (Model 737-14A) into a Teflon® bag. After the sample evaporated,

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approximately 1 mL of deionized (DI) water was added to the bag and allowed to partially

Total Organic Carbon and Evaporation Rate

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evaporate in order to increase the relative humidity in the bag to above 90% and thereby reduce

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condensation of gas-phase species on the walls of the Teflon® bag. 21 TVC levels were

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approximately 60% higher under humid conditions in the bag than under dry conditions. The gas

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from the bag was then sampled by a hydrocarbon analyzer (Model 55i, Thermo Scientific). The

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detection limit of the hydrocarbon analyzer is 0.2 parts per million (ppm); thus, concentrations

105

below 5.3 mg C L-1 could not be detected using this method. This measurement process was

106

performed shortly after receiving the samples and approximately 1 year later when

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environmental chamber experiments took place; TVC in the samples decreased on average by

108

25% over the course of this year (all TVC measurements are reported in Table S1).

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The TVC evaporation rate from samples was determined by passing 1 L min-1 of clean air over

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10 mL of a wastewater sample which had an exposed surface of 4.9 cm2. A 5 L Teflon® bag was



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filled and 50 µL of DI water were added to the bag and evaporated to increase relative humidity

112

to approximately 40% at room temperature – the Teflon® bag was agitated to dislodge small

113

droplets from the Teflon® surface and speed the evaporation process. The hydrocarbon analyzer

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was then used to determine the amount of evaporated carbon in ppm, which was converted to an

115

emission rate (mgC (L-m2-min) -1) assuming a constant emission rate over the course of this 5-

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minute measurement.

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2.2

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Experiments were performed in the Atmospheric Physicochemical Processes Laboratory

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Experiments (APPLE) chamber located at the University of Texas at Austin (UT-Austin). The

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APPLE chamber is a ~12 m3 Teflon ® bag suspended inside of a temperature-controlled room

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and has been described previously.22 The walls of the room are lined with UV lights which can

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be used to induce photolysis reactions. Before each experiment the bag was flushed for at least

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12 hours with clean air at a flow rate exceeding 100 L min-1. Following the injection of flowback

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wastewater (described below), polydisperse ammonium sulfate ((NH4)2SO4) seed particles were

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generated from a dilute aqueous ((NH4)2SO4) solution (Fisher Scientific, 99.5%) using an

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Aerosol Generation System (Brechtel, AGS Model 9200) and introduced to the chamber.

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Inorganic seed particles can be used to monitor wall loss rates 23 and to limit vapor wall losses to

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Teflon® walls. 24 Nitrous acid (HONO) was used as a source of OH and NOx in most

129

experiments. High NOx conditions were used in all experiments. HONO was created by slowly

130

adding 20 mL of 0.05 M H2SO4 to 10 mL of 0.1 M NaNO2 and then injected into the chamber by

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passing air over the solution. Photo-oxidation was initiated by turning on the UV lights, resulting

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in the photolysis of HONO and formation of OH, and reactions were allowed to proceed for at

Environmental Chamber Experiments


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least 4 hours with continuous UV light. Experiments were conducted in batch mode with no

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injections or dilution after the experiment was started.

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Two different injection methods were used to allow flowback samples to evaporate into the

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chamber. In method 1, 2-4 mL of sample were placed directly into the dry chamber at 30°C and

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allowed to completely evaporate over 16 hours. The chamber was then humidified to 25-35%

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RH to be more representative of ambient conditions and to minimize the amount of hydrocarbons

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on the walls of the chamber by passing air through heated water and into the chamber. In method

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2, 2-10 mL of sample was placed directly into the chamber at 20°C and allowed to evaporate for

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16 hours; 1-2 g of the sample evaporated during this time. Enough water to reach approximately

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25% RH was misted into the chamber and also allowed to evaporate during this time. Results

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from experiments utilizing method 1 are shown in Table 1 and are compared to results from

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experiments utilizing method 2 in Table 2.

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2.3

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The composition of PM1 (particulate matter smaller than 1 micrometer in diameter) formed in

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chamber experiments was measured using an Aerosol Chemical Speciation Monitor (ACSM)

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from Aerodyne Research Inc.25 In the ACSM, particles are flash-vaporized on a heater at 600 °C,

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and the resulting gas molecules are ionized using electron-impact ionization. This high-energy

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ionization method results in fragmentation of most molecules. The molecular fragments, which

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are measured by a quadrupole mass spectrometer, are attributed to four categories - organics,

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nitrate, sulfate, and ammonium - using a fragmentation table. 26 During this study the ACSM was

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operated at a time resolution (filter/sample cycle length) of approximately 90 seconds. For some

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experiments, the size distribution of particles was measured using a Scanning Electrical Mobility

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System (SEMS) from Brechtel Manufacturing, Inc. The SEMS uses a Differential Mobility

Instrumentation



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Analyzer (DMA) to size-select particles based on their electric mobility, which are then counted

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by a Condensation Particle Counter (CPC). The DMA continuously cycled between the voltages

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which select particles ranging from 5 to 1000 nm in diameter, resulting in a time resolution of the

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particle size distribution of approximately 60 seconds.

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Gas and particle-phase reaction products were monitored using a high resolution time of flight

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chemical ionization mass spectrometer (HR-ToF-CIMS 27,28) coupled to a filter inlet for gases

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and aerosols (FIGAERO 29) from Aerodyne Research, Inc. The HR-ToF-CIMS uses softer

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chemical ionization which results in minimal fragmentation of parent molecules. Hydronium

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water clusters (H3O+·(H2O)n) were used as reagent ions in this work. Hydronium water cluster

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ionization is most sensitive towards detection of moderately oxidized hydrocarbons; the ability to

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ionize and thus sensitivity is based on the relative proton affinity between the hydronium-water

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cluster and the parent molecule. 30

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Concentrations of NO and O3 were measured using Teledyne chemiluminescence NOx and

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absorption O3 monitors (200E and 400E, respectively); concentrations of NO2 were measured via

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an NO2 monitor from Environnement (Model AS32M), which uses a cavity attenuated phase

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shift (CAPS) method to directly measure NO2.31 The advantage of this direct NO2 measurement

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is that it does not rely on NO2 conversion to NO and therefore does not suffer from interference

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by other oxidized nitrogen compounds such as HONO and organic nitrates. 32 Volatile carbon

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emissions from samples of flowback wastewater were measured using a Thermo Scientific 55i

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direct methane and non-methane hydrocarbon analyzer. The 55i uses a column to separate

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methane and non-methane hydrocarbons which are quantified with a flame ionization detector

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(FID). The instrument uses a single FID response factor for all non-methane hydrocarbons and

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was calibrated using a mixture of methane and propane. Hydrocarbons dissolved in the flowback


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wastewater are expected to have a longer chain length and therefore a higher response factor in

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the FID. However, for this alkane-dominated mix of hydrocarbons the instrument’s response is

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expected to be proportional to carbon chain length, and TVC measured is reported in units of ppb

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C.

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Data Analysis

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Data from the ACSM were analyzed in Igor Pro (Wavemetrics) using the software package

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“ACSM Local,” which includes a correction for relative ion transmission efficiency as well as

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changes in the flow rate throughout the experiment. Standard fragmentation and batch tables

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were applied. 26 The ACSM does not detect all sampled particles, primarily due to particle

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bounce at the vaporizer, resulting in a collection efficiency (CE) smaller than 1. A composition

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dependent collection efficiency 33 was applied to correct for this. SEMS data were available for

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approximately half of the experiments conducted. For these experiments, the SEMS volume

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concentration was converted to mass concentration using the densities 1.77 g cm-3 for

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ammonium sulfate, 1.72 g cm-3 for ammonium nitrate, and 1.4 g cm-3 for organics. 34 The mass

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estimates from the SEMS were typically within 10% of the total mass estimated by the ACSM

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using the composition dependent collection efficiency; ACSM data are reported in Tables 1 and

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2. A characteristic time series of mass concentrations from the ACSM and SEMS (not corrected

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for wall losses) is shown in Figure 1a. A depositional wall loss correction 23 was applied. PM

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formation seemed to cease 20-30 minutes after the start of photo-oxidation (Figure 1b) and the

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total amount of PM formed was evaluated then when the effects of wall losses are minimal.

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The fraction of total emitted carbon in the particle phase after photo-oxidation was evaluated

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using estimates of the organic mass to organic carbon ratio (OM:OC) from the ACSM. Aiken et



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al. 35 showed that the fraction of the organic signal at m/z 44 (f44) correlates with the average

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elemental O:C ratio of organic aerosol components. The correlation was recently updated to 36:

203 204

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O:C = 4.31 × f44 + 0.079

(1)

Aiken et al. 35 also observed a correlation between O:C and OM:OC: OM:OC = 1.29 × O:C + 1.17

(2)

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Thus, for each experiment measurements of f44 were used to estimate the average OM:OC of

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organic aerosol components, and the estimated OM:OC was used together with ACSM OM

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measurements to calculate the total mass of organic carbon (OC).

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Data from the HR-ToF-CIMS were analyzed in Igor Pro using Tofware, the software provided

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with the instrument. The data were first mass calibrated based on HR-ToF-CIMS reagent ions

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and other known ions. The baseline was subtracted and the average peak shape was found so it

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could be used for high resolution analysis, through which multiple ions can be identified at any

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given integer m/z. Ions up to m/z 350 were analyzed in high resolution mode. After ions were

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identified in the high resolution spectrum, the peaks were integrated to yield the concentration

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time series of each ion. Analyte ion concentrations were then normalized by the reagent ion

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concentrations: the sum of H3O+, H3O+·(H2O) and H3O+·(H2O)2. This correction accounts for

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changes in reagent ion concentrations and instrument sensitivity during and between

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experiments. Relative humidity can affect instrument sensitivity, but this was set to

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approximately 30% for all experiments. During gas-phase sampling particles were collected on a

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Teflon® filter in the FIGAERO. After approximately 30 minutes of collection a heated (125°C)

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stream of clean, dry air was used to desorb the organic aerosol compounds into the gas phase so

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they could be measured by the HR-ToF-CIMS.


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2.4 Box modeling

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Box-modeling was conducted in the SAPRC framework 37 in order to characterize the oxidative

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environment and the effects of initial NOx concentrations on formation of HNO3. Experiment 9

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was chose as representative experiment and modeled with the gas-phase chemical mechanism

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CB6r5.38 Measured concentrations of hydrocarbons are used and represented as hexane and

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heptane (SAPRC model species: PAR). Results from chamber characterization experiments are

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utilized to account for chamber wall effects such as the chamber radical source, O3 decays, and

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NOx offgasing.39 The model was initiated with measured concentrations of nitrogen oxides (NO,

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NO2) and estimated concentrations of HONO. HONO was not measured directly but was

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estimated from the difference in initial NO2 measurements reported by the chemiluminescence

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NOx monitor (where HONO is measured as NO2) and CAPS NO2 monitor (which does not detect

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HONO). The sensitivity of the chemiluminescent NOx monitor to HONO has not been measured.

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A 20% uncertainty was assumed and initial HONO concentration in the model was varied within

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20% to minimize differences in measured and modeled concentrations of NO and NO2.

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3

Results and Discussion

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3.1

Volatile carbon content and emission rate

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Experimental results are summarized in Table 1, including measured levels of total volatile

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carbon (TVC) and its emission rate. The pH of all samples was measured and lied in the region

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7-8. Losses of organic vapors to the walls of the Teflon® bag and the instrument during these

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measurements could cause an underestimation of TVC. The increased RH during sampling was

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used to mitigate this effect, but reported measurements should still be considered a lower-bound

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estimate of TVC. According to Field et al.40, only hydrocarbons containing up to 11 carbon



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atoms are efficiently measured by the hydrocarbon analyzer; larger molecules are lost to the

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walls of the instrument. The data in Table 1 were collected at the time of the chamber

247

experiments, approximately 1 year after the samples were collected.

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Measured levels of TVC ranged from below the detection limit (5.3 mg L-1) to 92 mg L-1 with an

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average of 29 mg L-1. Some of the measured TVC levels are comparable with measurements of

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total dissolved organic carbon (TOC) in produced water in the Permian Basin from Khan et al.41,

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which ranged from 63 to 146 mg L-1 (average 109 mg L-1). Other samples (e.g. samples 3,5,6)

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had significantly lower levels of TVC. Levels of TOC in natural surface waters in the Permian

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Basin are typically 0-10 mgC L-1. 42 Measurements of TOC differ from measurements of TVC as

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TOC also includes organic compounds that do not evaporate. Thus, TOC measurements are

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expected to be higher than TVC, and most flowback wastewater samples contain more carbon

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than natural surface waters. 42

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Several factors can affect the amount of carbon in wastewater samples. There is variability in the

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carbon content of the produced water based on geology of the formation and age of the field. 43

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The organic carbon content of fracturing fluid can also vary significantly as it is designed to have

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the viscosity, friction, and other characteristics best suited to fracture a particular geologic

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formation. 7 Hayes 44 reported that for sites in the Marcellus shale gas region TOC in fracturing

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fluid before injection was 5.6 – 1260 mg L-1 (median 226 mg L-1) and then decreased to 3.7 –

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388 mg L-1 (median 62.8 mg L-1) and then 1.2 – 509 mg L-1 (median 38.7 mg L-1) after 5 and 14

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days of flowback, respectively. Thus, while median concentrations of TOC decreased over time,

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TOC increased in some samples (maximum 388 mg L-1 after 5 days and maximum 509 mg L-1

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after 14 days). This suggests a difference in the relative contribution of fracturing fluid and

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produced water to TOC in flowback wastewater. This work showed no clear trend between the


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TVC content of samples and the number of days after fracturing when the sample was collected.

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The carbon content of wastewater is also affected by varying efficiency in separation processes

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which are unknown for the samples collected here.

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Hydrocarbons evaporating from these samples were not detectable with the HR-ToF-CIMS using

272

H3O+(H2O)n ionization. This suggests that most evaporated hydrocarbons are nonpolar, and that

273

polar compounds are below the detection limit of the HR-ToF-CIMS using H3O+(H2O)n

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ionization. This is consistent with the observations of Tellez et al. 45 and Khan et al. 41that

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alkanes are the dominant hydrocarbon in produced water with straight chain alkanes the most

276

abundant 41.

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The evaporation rate of volatile carbon ranged from 122 to 5700 mgC (L-m2-min) -1 (average

278

1357 mgC (L-m2-min) -1) and correlated with TVC content for most samples (Table 1). The

279

exceptions were samples 4 and 9 which had very high evaporation rates. This may be an

280

indication that these samples contain a higher fraction of lighter hydrocarbons compared to the

281

other samples; lighter hydrocarbons have a higher vapor pressure distribution and evaporate

282

more quickly. The significant correlation between evaporation rate and TVC (R2 = 0.61, or R2 =

283

0.82 excluding samples 4 and 9) suggests that samples are composed of organic compounds with

284

similar vapor pressure distributions despite having very different concentrations of these organic

285

compounds.

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3.2

Particulate Matter Formation

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Particulate matter formed as evaporated flowback wastewater was exposed to a high-NOx photo-

289

oxidative atmosphere. PM data from a characteristic experiment is shown in Figure 1 (Expt 4).



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Organic compounds formed organic aerosol after oxidation, and ammonia emissions from the

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wastewater combined with HNO3 produced in the high NOx environment to form ammonium

292

nitrate. The amount of particulate ammonium attributable to the ammonium sulfate seed particles

293

was calculated from the ammonium:sulfate molar ratio, and extra ammonium was attributed to

294

ammonium nitrate. The photo-oxidation of hydrocarbons under high-NOx conditions is also

295

expected to form organic nitrates, as confirmed by data from the FIGAERO-CIMS (see section

296

3.2.1). Mass spectrometer data from the ACSM can and have been used to distinguish between

297

organic and inorganic nitrate.46 This is not possible for the data presented here due to the much

298

higher concentrations of inorganic nitrate compared to the concentration of organic (nitrate) in

299

the aerosol and uncertainties regarding the fragmentation pattern of organic nitrates from

300

different precursors. 46 The analysis presented here thus allows for calculation of two bulk PM

301

components: organic PM and ammonium nitrate, as illustrated in Figure 1b (corrected for wall

302

loss and collection efficiency, see Section 2.3). HONO-initiated photochemistry proceeds

303

quickly and little increase in PM mass concentrations was observed after the first 20 minutes of

304

photo-oxidation. At this point 20 minutes of data were averaged and summarized in Table 1 as

305

the amount of organic PM formed per mL of evaporated flowback wastewater.

306

307

3.2.1

Organic PM

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Organic PM formation correlated linearly with the measured amount of TVC (R2=0.85). This

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suggests that though concentrations are different, the carbon compounds that compose the TVC

310

in different samples have similar potential to form organic aerosol.


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The compounds formed from the oxidation of evaporated flowback wastewater can be observed

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with the HR-ToF-CIMS. With the FIGAERO, oxidized particle-phase compounds can also be

313

observed. Figure 2a shows time series of the 20 gas phase compounds which exhibited the

314

highest signal in the HR-ToF-CIMS during Expt. 1. Most of these compounds have 5-10 carbon

315

backbones; some have more than 10 carbon atoms – a table of the molecular formulae of the ions

316

summarized in Figures 2a and 2b and their double bond equivalency is provided and discussed in

317

the supplemental information (Table S2). The vertical axes in Figure 2 are arbitrary as explicit

318

calibration for all of these compounds is not possible, especially as only empirical formulae are

319

known and not the molecular structures. Figure 2b shows the 20 particle-phase compounds

320

which had the highest signal during a heated filter desorption during Expt.1. Compounds with

321

11-19 carbon backbones dominate the signal, with the highest contribution from molecules with

322

14-16 carbon atoms. The composition of organic vapors evaporating from the wastewater is

323

expected to be dominated by molecules of carbon number less than 10. Thus, the prevalence of

324

much larger molecules in the particle phase suggests that oligomerization reactions are an

325

important pathway for the formation of organic aerosol from evaporated flowback wastewater

326

and emphasizes the importance of atmospheric processing in the formation of OA. Interestingly,

327

of the twenty particle-phase compounds with the highest signal, half appear to be organic

328

nitrates. In the gas phase no organic nitrates were among the most abundant 20 compounds. This

329

analysis was also performed for Expt. 9 and the results were similar. It appears that organic

330

nitrates play an important role in the formation of particulate matter but have a smaller role in

331

compounds which remain in the gas phase. The addition of the –ONO2 functional group lowers

332

the vapor pressure of a compound by approximately two orders of magnitude, similar to an –OH

333

functional group 47, making partitioning to the particle phase more likely. Organic nitrates are



334

expected to be present in the gas phase; however, they are not among the 20 most abundant

335

species and therefore do not appear on Figure 2a.

336

The ratio of total aerosol carbon, calculated from organic mass measured by the ACSM, to

337

measured TVC is shown in Table 1 (particle carbon fraction). These estimates indicate that

338

between 11 and 61% (average 33%) of the emitted carbon partitions to the particle phase. Thus,

339

the formation of particulate matter is an important fate of hydrocarbons emitted from flowback

340

wastewater. A calculation of SOA mass yield would require determination of all hydrocarbons’

341

identity, as well as their initial and final concentrations, which are not available for these

342

experiments.

343

The two methods of flowback wastewater evaporation for chamber experiments provide insight

344

into emissions and potential for particulate matter formation (see Table 2). In method 2, more

345

flowback wastewater was placed into the chamber but did not evaporate completely over the 16

346

hours allotted. Despite this and the lower evaporation temperature (20°C instead of 30°C), OA

347

formation/mL (normalized by mL placed in the chamber, not by the amount evaporated) was

348

similar to that in method 1 even though only 20% of the water evaporated. This suggests that

349

hydrocarbons evaporate much more quickly than the water in which they are dissolved.

350

3.2.2

351

Table 1 also summarizes measurements of particulate ammonium nitrate in units of micrograms

352

formed per mL of flowback wastewater evaporated. Because experiments were conducted under

353

high-NOx conditions, an excess of HNO3 was available and ammonia was the limiting reactant in

354

ammonium nitrate formation in these experiments. Thus the numbers in Table 1 can be

355

considered an upper limit to the potential for ammonium nitrate formation from evaporated

Ammonium Nitrate


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356

flowback wastewater. Ammonium nitrate accounted for most of the total PM formed in chamber

357

experiments (92% on average).

358

The concentrations of NOx used in these experiments were greater than what would be expected

359

in ambient environments (mostly due to the use of HONO) as summarized in Table S3.The

360

SAPRC model was used to evaluate the effects of initial NOx levels on ammonium nitrate

361

formation by decreasing the levels of NO, NO2 and HONO in the model, letting the model

362

calculate the amount of HNO3 formed and calculating the amount of ammonium nitrate formed

363

from thermodynamic equilibrium, measured ammonium concentration and temperature. In

364

modeling Expt. 9, decreasing NO, NO2 and HONO by 90% (from 500 ppb to 50 ppb) results in

365

predicted HNO3 of 46 ppb at 300K (compared to the original 142 ppb) and predicted formation

366

of ammonium nitrate of 353 µg/ml (compared to the measured 378 µg/ml). Thus, lowering the

367

concentrations of NOx and HONO by 90% lowers the predicted formation of ammonium nitrate

368

by less than 10%, and ammonium nitrate formation is still high under much more moderate NOx

369

conditions. In addition, lowering the temperature to 285K increases ammonium nitrate formation

370

to 385 µg/ml under the more moderate NOx conditions.

371

While OA formation was similar between injection methods 1 and 2, ammonium nitrate

372

formation appears to be higher when the sample evaporates completely. This implies ammonia

373

evaporates more slowly than organic compounds but faster than water.

374

3.3

375

Ammonia in water can be in the form of dissolved NH3 or aqueous NH4+, depending on the pH

376

and temperature, so for simplicity it is usually measured as total ammonia nitrogen (TAN). In

377

these experiments TAN concentrations are estimated by assuming that all ammonia evaporated

Ammonia Emissions



378

from samples forms ammonium nitrate (see Table 1). At these high-NOx conditions partitioning

379

estimates indicate that over 95% of ammonia partitions to the particle phase as ammonium

380

nitrate, so for simplicity gas-phase ammonia is neglected. Estimated TAN concentrations in these

381

samples ranged from 21 to 82 mg L-1 (average 59 mg L-1). All estimated TAN levels were above

382

the acute EPA criterion (17 mg L-1) and more than 10 times the chronic EPA criterion (1.9 mg L-

383

1

384

The source of ammonia in this wastewater is unclear as it can be a contaminant of produced

385

water 49 and a fracturing fluid additive. 50 In the Marcellus Shale, Hayes observed TAN

386

concentrations of 0.28 – 441 mg L-1 (median 5.9) in freshly prepared fracturing fluid which

387

changed to 29.4 – 199 mg L-1 (median 71.2) and then 3.7 – 359 mg L-1 (median 124.5 mg L-1)

388

after 5 and 14 days of flowback, respectively. 44 The increase in median TAN concentrations

389

with time suggests that produced water can be a source of ammonia. However, the Permian

390

Basin, where the samples analyzed here were collected, is known for having lower ammonia

391

levels. 51 The data from Hayes also suggest that ammonia can be a component of the original

392

fracturing fluid, and ammonium salts are a common ingredient of fracturing fluid. 7,50 Thus, the

393

original fracturing fluid is likely the largest source of ammonia for these samples from the

394

Permian basin.

) for natural fresh waters. 48

395

396

3.4

Implications

397

A simplified yet illustrative calculation can demonstrate the potential effect of PM formation

398

from evaporated flowback wastewater on regional PM2.5, using the State of Texas as example. In

399

2008 an estimated 40 million m3 of water were used for hydraulic fracturing across Texas. 17 We


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400

assume that 50-150% of this water flows back to the surface 18, and that evaporated water has the

401

potential to form ammonium nitrate and OA measured in this work (Table 1). Estimating

402

evaporation of volatile compounds from vented storage tanks is virtually impossible considering

403

very limited data on wastewater composition and the storage tanks used. Assuming that 10% of

404

volatile PM precursors evaporate, subsequent atmospheric transformations would produce 578-

405

1735 tons/year of ammonium nitrate and 53-160 tons/year of OA (total PM production of 631-

406

1895 tons/year); complete evaporation of PM precursors would produce ten times as much PM..

407

For comparison of this potential PM production, in the state of Texas rig emissions of PM2.5

408

(from diesel engines) were estimated at 2470 tons/year in 2008. 52 This suggests that formation

409

of PM2.5 from evaporated wastewater can significantly add to the total PM2.5 burden from oil and

410

gas production activity.

411

As for ammonia emissions, the 2014 EPA National Emissions Inventory 53 for point, onroad,

412

non-point and nonroad sources for the Delaware sub-basin counties (the origin of samples used

413

in this work) estimates ammonia emissions as roughly 3,100 tons per year. Using a produced

414

water volume of about 58 billion gallons for the period 2005-2015 18, estimated emissions of

415

ammonia from flowback wastewater (using average TAN of 59 mg L-1 measured in this work

416

and assuming complete evaporation) is 1298 tons/year, over 40% of the total ammonia emissions

417

estimated from the latest inventory. Thus, evaporation from flowback wastewater can

418

significantly contribute to ambient ammonia concentrations.

419

Under more moderate NOx conditions the atmospheric processing of evaporated wastewater can

420

also result in significant formation of ozone. While not the focus of this work, in one experiment

421

in which hydrogen peroxide and NO were used as sources of ·OH and NOx (approximately 1:1

422

VOC:NOx), ozone production was observed. In all other experiments, higher NOx concentrations



423

suppressed the accumulation of ozone. Thus, evaporation of hydraulic fracturing wastewater and

424

its atmospheric processing can increase concentrations of ozone and particulate matter. The

425

organic compounds which evaporate from the wastewater appear to have high carbon number

426

and form OA at high efficiency. Ammonia emissions appear to be significant and form

427

particulate ammonium nitrate in high NOx environments. This is especially important to the total

428

PM2.5 burden as 1 kg of ammonia can be transformed into up to 4.7 kg of ammonium nitrate.

429

The potential of flowback wastewater to influence air quality will depend on the extent of

430

evaporation. Quantifying evaporation of flowback wastewater requires several pieces of

431

information which are currently largely unavailable. This includes the storage method

432

(evaporation pond or ventilated storage tank) and the extent of evaporation or venting from

433

different storage methods (e.g. how tanks are ventilated and whether the ventilation process is

434

working as expected). In addition, evaporation rates of organic and ammonia compounds from a

435

complex mixture of flowback wastewater need to be quantified, which is complicated by the fact

436

that the wastewater composition is generally unknown. Meteorological conditions will also affect

437

evaporation, including the ambient temperature (which will affect tank temperature, which can

438

and should be measured) and the ambient wind speed. This work shows that wastewater

439

evaporation can be a significant source of PM and motivates quantification of evaporation from

440

flowback wastewater, which could be used in air quality models together with results from this

441

work and future experiments to assess the regional air quality impacts of flowback wastewater

442

management strategies.

443

4. Acknowledgements

444

Funding was provided by start-up support from the University of Texas at Austin.


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445

Supporting Information

446

Additional details on sample origin, collection and storage (S1), TVC measurement comparison,

447

1 year apart (S2), Molecular formulae of most abundant ions identified in gas and particle phases

448

(S3), NOx and HONO concentrations at the beginning and end of each experiment (S4).

449



Table 1. Flowback wastewater properties and mass of particulate matter formed

450

OA formeda (µg mL1 )

PM C Fraction

277

41.2

0.34

L1W3S

21

92

10.2

0.21

L2W3T1

122

67

300

4.5

N/A

L2W3T2

92

5286

66

294

91.2

0.53

L2W2S

5

Formation of Particulate Matter from the Oxidation of Evaporated

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