Small Volume Solid Phase Extraction Method for Comprehensive Analysis of Neonicotinoids, Their Metabolites, and Related Pesticides in Water

Abstract

Neonicotinoid insecticides (NEOs) such as clothianidin, imidacloprid, and thiamethoxam are used worldwide. The occurrence of their degradates, for instance, clothianidin-n-desmethyl (CLO-N-DES), clothianidin-urea (CLO-U), imidacloprid urea (IMI-U) and olefin (IMI-O), as well as thiamethoxam urea (THX-U), have seldom been documented in water due to the lack of a sensitive analytical method. In this study, a method only requiring 12 mL of water sample was developed and validated to quantify 8 NEOs, 13 metabolites, and 3 related insecticides using solid phase extraction (SPE) coupled with HPLC-MS/MS. The method demonstrated good linearity (r²>0.99), with limits of detection (LOD) ranging from 0.16–1.21 ng/L and limits of quantification (LOQ) from 0.54–4.03 ng/L in water samples. Validation showed accuracy between 70–130% and precision below 15% for most analytes. The method’s performance was comparable to, or better than, existing methods, with the advantage of requiring much smaller sample volumes. Using this method, we monitored the occurrence and seasonal variability of NEOs and their metabolites in various surface water and groundwaters matrices from across Iowa. For example, analysis of water samples from private wells across three Iowa counties detected several NEOs, with notable findings including the first detection of flupyradifurone (FLU) in Iowa well water. Surface water analysis from five locations revealed frequent detection of NEOs and their metabolites, with some concentrations exceeding U.S. EPA chronic toxicity benchmarks for freshwater invertebrates. In addition, this is the first study to demonstrate the occurrence of CLO-N-DES, CLO-U, and THX-U in US surface water. The study helps advance analytical methods for NEOs and their metabolites while also highlighting their widespread occurrence in Iowa waters and associated ecological risks, emphasizing the need for more comprehensive monitoring of these compounds.

1. Introduction

Neonicotinoids (NEO) are a class of insecticides used in agriculture for protection against chewing insects such as plant hoppers and thrips^[1] and for veterinary purposes.^{[2, 3]} Although the use of NEOs is banned in some countries.,^[4] their global use continues to grow. NEOs are registered in over 120 countries^[1] and account for nearly one-third of all pesticides traded globally.^[5] In 2023, NEOs had a total market value of $5.5 billion USD.^[6] This is expected to grow to $8.6 billion USD by 2030.^[6] In the United States (US), NEOs are applied annually on up to 140 different crops, with the three most used NEOs being imidacloprid (IMI), clothianidin (CLO), and thiamethoxam (THX).^[1]

Because of their widespread agricultural use, neonicotinoids are frequently detected in drinking water^[7–13], surface water^[14–19], and groundwaters.^{[11, 12, 20–22]} Their occurrence has been well documented in the waters of the “corn belt” states of Iowa, Minnesota, and Wisconsin.^{[11–13, 18–30]} For example, across these three midwestern US states, the concentrations of individual NEOs have been detected as high as 3,500 ng L⁻¹ in surface waters.^[18] In Iowa, maximum concentrations for CLO, IMI, and THX in drinking water across both municipal drinking water systems and private wells have been reported to be 140.5 ng/L, 29.3 ng/L, and 13.1 ng/L, respectively.^{[8, 13, 21]}

Their widespread occurrence in the environment has raised concerns about potential adverse effects on biodiversity.^[31–33] NEOs exposure is thought to exert direct adverse impacts on terrestrial and aquatic invertebrates^{[14, 33–37]} and vertebrates^{[32, 38, 39]} and indirect effects via trophic cascades that can change food webs.^{[33, 40–42]} In response to these concerns, the U.S. Environmental Protection Agency (EPA) has established Aquatic Life Benchmarks (ALB) for chronic and acute toxicity toward freshwater species for most NEOs (Table S1).^[43]

Neonicotinoids have also been shown to breakdown in the environment.^[3] Their metabolites and other transformation products form through different processes like microbial^[44–47] and fungal^{[48, 49]} degradation, photolysis^[50–55], and hydrolysis.^{[8, 51, 56–59]} Research has also shown that transformation products with potentially greater mammalian toxicity can be formed during water treatment when the insecticides react with free chlorine,^[56] which is commonly used as a chemical disinfectant. Nevertheless, most occurrence studies looking at groundwater and surface water, especially in the US, have focused on the primary parent NEOs. Accordingly, data on the occurrence of known NEO metabolites in drinking water, surface water, and groundwater is still limited, which leaves gaps in our understanding of their fate and effects in the environment.

There is also evidence to suggest that drinking water may represent an important source of human exposure to NEOs and their metabolites, highlighting the need for more comprehensive monitoring. Analysis of human biospecimens in Iowa has suggested broad exposure to different NEOs beyond what has previously been measured in the environment.^[13] A median of 10 different NEOs and/or metabolites were detected in urine, with clothianidin, nitenpyram, thiamethoxam, 6-chloronicotinic acid, and thiacloprid amide detected in every urine samples analyzed.^[13] Dinotefuran, imidaclothiz, acetamiprid-N-desmethyl, and N-desmethyl thiamethoxam were found in ≥70% of urine samples.^[13] The study showed that consumption of drinking water was an important human exposure pathway, but given the absence of more comprehensive testing of drinking water resources, the researchers were left to conclude that diet and/or other exposure pathways (e.g., occupational, house dust) may better explain the diverse exposure profile than water contamination.^[13]

Given the threats to biodiversity and the potential for human exposure, further environmental monitoring of NEOs, their known metabolites, and transformation products is warranted. Most studies documenting their occurrence in water have used solid phase extraction (SPE) methods that while sensitive can also have significant limitations. Most SPE methods achieve their sensitivity by processing large volumes of water, as much as 1 L. As a result, processing samples can be time consuming, and large sample volumes make it difficult to collect, transport, and store the large number of samples necessary to regularly monitor the occurrence and fate of NEOs and their metabolites across various water resources. Other methods using direct aqueous injection (DAI) are more efficient and require smaller volumes, but they lack the sensitivity of SPE and thus may be unable to detect the presence of these chemicals in most samples. A sensitive, small volume method is needed to improve the ease and frequency of routine testing of diverse water matrices for NEOs and their degradates.

Herein we developed a sensitive, small volume method for 24 analytes, including 8 NEOs, 13 metabolites, and 3 related insecticides. The NEOs and metabolites analyzed included acetamiprid (ACE), acetamiprid-n-desmethyl (ACE-N-DES), CLO, clothianidin-desmethyl (CLO-N-DES), dinotefuran (DIN), dinotefuran-n-desmethyl (DIN-N-DES), IMI, desnitro-imidacloprid hydrochloride (D-IMI), 5-hydroxy-imidacloprid (5-OH-IMI), imidacloprid olefin (IMI-O), imidacloprid urea (IMI-U), imidaclothiz (IMZ), thiacloprid (THC), thiacloprid-amide (THC-A), THX, thiamethoxam-n-desmethyl (THX-N-DES), clothianidin urea (or thiamethoxam metabolite CGA 353968, CLO-U), thiamethoxam urea (or thiamethoxam metabolite CGA 355190, THX-U), nitenpyram (NIT), and nitenpyram-n-desmethyl (NIT-N-DES). Flupyradifurone (FLU) – a butenolide, flonicamid (FLO) - a pyridine, sulfoxaflor (SUL)- a sulfoximine, and the sulfoxaflor metabolite X11719474 (SUL-X). Inclusion of these analytes into the developed method were motivated by several factors including historical use in Iowa based upon USGS pesticide use estimates (SUL, FLO) ^[30], veterinary use (NIT)^[60], documented human exposure (IMZ) ^[13], and/or because they are considered as an alternative to NEOs (FLU) and have been used on NEO resistant pests (FLU).^[61]

The three primary objectives of this study were to (1) develop a broad-spectrum method including many NEOs, known metabolites, and related pesticides with comparable sensitivity to large volume SPE but with a sample volume comparable to DAI, (2) validate the method’s efficacy across different types of water samples relevant to NEO transport and fate in the environment, and (3) use the validated method, to characterize the occurrence of these chemicals in various water matrices across Iowa. Occurrence was assessed from planting to harvest seasons and in various types of water samples (e.g., private well water, wastewater effluent discharge, surface water) to determine overlooked sources for their entry into the environment, establish their abundance and persistence in different water resources over time, and assess associated ecological risks and potential routes of human exposure.

2. Materials and methods

2.1. Sample Collection

Environmental monitoring was conducted across Iowa over a 10-month period from March 2023 to January 2024. During this time, periodic grab samples were collected from various water resources in Iowa. These resources included a municipal effluent-impacted stream, agriculturally impacted streams, and private drinking water wells.

Cardinal Creek, previously known as Sewer Creek, is situated in Newton, Iowa, and it receives treated effluent from Newton’s municipal wastewater treatment facility. Samples were taken from the effluent leaving the Newton Water Pollution Control facility and from Cardinal Creek upstream and downstream of the effluent discharge point. The Newton Water Pollution Control Plant operates as a trickling filter/activated sludge facility with UV disinfection, typically producing more than 3 million gallons (or 11,000 metric tons) per day of treated effluent. Preliminary treatment includes two mechanically cleaned screens and two grit removal systems. Primary treatment utilizes three primary clarifiers to remove suspended solids. For secondary treatment, the trickling filters primarily target organic load (i.e., BOD₅) remaining after primary clarification, whereas the activated sludge process focuses on ammonia oxidation. The disinfection system employs a Trojan UV Signa system, with the disinfection season running from March 15 to November 15 annually.

Burr Oak Creek (near Perkins, IA), Pilot Creek (near Rolfe, IA), and the Little Floyd River (near Sanborn, IA) are locations of United States Geological gages (stations 06483495, 05476735, and 06600030, respectively). The sampling location in Burr Oak Creek is situated in an agriculturally dominated portion of the watershed, but several miles upstream it receives effluent discharged from a wastewater treatment lagoon for the city of Hull, IA. The sampling locations in Pilot Creek and the Little Floyd River were both within agriculturally dominated watersheds with no known municipal effluent discharges upstream.

Surface water samples were also collected from the Mississippi River at Pleasant Valley, Iowa, from the shore near three residences from which private well water (depth range: 45 – 60 m) was also sampled. Additional private well water samples were collected from three, shallow private drinking water wells (depth range: 7 – 18 m) in O’Brien County, Iowa, in close proximity to the gauge station for the Little Floyd, and two shallow private wells (depths = 4.5 m) in in Johnson County, Iowa in the city of Hills.

Water samples from these locations were collected for their diversity in both bulk water quality and likelihood of neonicotinoid occurrence. All grab samples (40 mL) were collected in amber glass bottles (Fisher Scientific, Pittsburg, Pennsylvania, USA) without filtering. After collection, samples were transported on ice to the laboratory. Upon arrival, the sample pH was measured, resulting in a range of 7.6 ± 0.5 across all sample types. Samples were then stored at 2–6 °C in the laboratory prior to analysis.

2.2. Preparation of native standard solutions, internal standard spiking solution, and calibrators

Reagents and chemicals can be found in the supporting information. For standards purchased as solid, individual stock solutions were first prepared at a concentration of 1 mg/mL by exactly weighing the solid compound and then dissolving in acetonitrile. Next, a 1000 ng/mL mixed stock solution was prepared every twelve months by combining individual analyte solutions and then diluting with acetonitrile. To prepare the spiking solutions, the concentrated stock was then further diluted to 100-, 10-, and 1-ng/mL solutions with acetonitrile and replaced every three months. All standard solutions were stored at <−10 °C. To prepare calibrators, 1000 mL calibration standards of concentrations ranging from 0.01 to 25 ng/mL were fresh prepared by spiking the appropriate volume of either a 100-, 10-, or 1 ng/mL analyte standard solutions into 0.25% formic acid in acetonitrile/Milli-Q water (1:9 v: v) before each run. The standard concentration and volume used for calibrators are summarized in Table S3.

Each stable isotopically labeled analog in solid form was first prepared at 0.2–0.5 mg/mL by dissolving in acetonitrile. The mix ISTD stock solution was prepared by combining all individual solutions and then diluted with acetonitrile, reaching a concentration of 1000 ng/mL. The ISTD spiking solutions were prepared by further diluting the concentrated solution to 100- and 10- ng/mL with acetonitrile. 10 mL of 100 ng/mL ISTD spiking solution was added to all calibrators, resulting in a concentration of 1 ng/mL for all isotopically labeled analogs.

2.3. Quality control and unknown sample preparation

For both unknown and quality control (QC) samples, 12 mL of water was first transferred to a disposable borosilicate 16×100 mm glass tube, and 10 mL of 10 ng/mL ISTD solution was spiked into each sample. Analyte stock solutions were only spiked into QCs (Table S3). The samples were then extracted by SPE automatically using a Vivace^Tm Duo cleanup station (PromoChrom Technologies Ltd., Richmond, British Columbia, Canada). Strata^™-X-CW 33 μm Polymeric Weak Cation cartridge tubes (60 mg/3mL; Part No: 8B-S035-UBJ, Phenomenex, Torrance, California, USA) were used during extraction and first washed with 5 mL methanol, followed by a wash with 5 mL Milli-Q water. Water samples were then loaded onto the cartridge. After sample loading, the sample tube and SPE cartridge was rinsed with 3 mL of Milli-Q water, followed by drying under vacuum for 30 seconds after air purge. Target analytes were eluted with 6 mL of 0.25% formic acid in acetonitrile/ ethyl acetate (8:2 v: v). Next, the elute solvent was evaporated to dryness using a Biotage TurboVap LV (Uppsala, Sweden). Extracts were then reconstituted with 100 mL of 0.25% formic acid in acetonitrile/Milli-Q water (1:9 v: v) mixture and transferred into an LC vial with insert for analysis. The SPE procedure for Vivace^Tm Duo cleanup station is described in Table S4.

2.4. Instrument analysis

The HPLC tandem mass spectrometry analytical system includes an Agilent 1290 module coupled to an AB Sciex QTRAP 6500 plus triple quadrupole mass spectrometer (AB Sciex, Framingham, Massachusetts, US) using a Kinetex Biphenyl column (2.1 × 100 mm, 5 μm; Phenomenex, Torrance, California, USA) for separation. Mobile phases were 0.1% formic acid in Milli-Q water (A) and 0.1% formic acid in acetonitrile (B). The LC flow rate and mobile phase gradients are summarized in Table S5. This LC condition was used to achieve high selectivity, short retention time (6 min), and good peak shape. An example of LC-MS/MS chromatogram for 2.5 ng/mL of analytes spiked in solvent is shown in Figure S1. The MS/MS was conducted using an electrospray ionization (ESI) source in the positive ion mode with scheduled multiple reactions monitoring (SMRM) for all analytes except for imidacloprid olefin and its isotopically labeled standards, which were conducted in negative mode. Curtain gas, CAD gas, ion source gas 1, ion source gas 2, and the source temperature were set at 30 psi, 9 psi, 50 psi, 50 psi, and 550 °C, respectively. IonSpray voltage were set at 5500 for the positive mode and −4500 V for the negative mode. An MRM detection window of 30 s and target cycle time of 750 ms were used for the analysis. Entrance potential was kept at 10 V. The exact masses to monitor, declustering potential (DP), collision energy (CE), and collision cell exit potential (CXP) were optimized for each ion transition by infusing a test solution of the analyte. All results are summarized in Table S6.

2.5. Method validation

A challenge with matrix selection for method validation is the widespread use of NEOs, particularly imidacloprid, clothianidin and thiamethoxam. Preliminary analysis of different matrices investigated herein including surface water, shallow well water, and wastewater effluent routinely detected appreciable levels of analytes, resulting in high background signals for certain analytes and confounding the reliable use of these matrices in our method validation approach. Accordingly, we elected to use deep well water (depth = 472 m) for method validation because it was free of all analytes and had similar pH value and ionic content to the other water matrices we investigated.

Quality control samples were prepared by spiking the well water with native target analytes to create a matrix blank, and samples with analytes at four concentration levels: QC Low (QCL) (0.42 ng/L in the water sample, 0.05 ng/mL in the extract), QC Medium Low (QCML) (4.17 ng/L in the water sample, 0.5 ng/mL in the extract), QC Medium High (QCMH) (10.42 ng/L in the water sample, 1.25 ng/mL in the extract), and QC High (QCH) (25 ng/L in the water sample, 3 ng/mL in the extract). All QCs were prepared and then extracted as described in section 2.4.

Method performance was evaluated by determining method linearity, sensitivity, specificity, accuracy, and precision. Linearity was evaluated by performing quadratic regressions over several 12-point calibration curves and analyzing residual plots (r² ≥ 0.98). Weighting with 1/x was used. An error of 15% from the nominal value was set as the acceptable limit, except for the lowest calibrator, which was allowed to be 30%. The limit of detection (LOD) and quantification (LOQ) were calculated by repeatedly measuring QCL to assess method sensitivity. For this, 18 measurements from 6 QCL samples (3 injections per sample) were performed on different days. The standard deviation (S₀) of the results was determined, and LOD and LOQ were calculated as 3- and 10-times S₀, respectively.

The method selectivity for each analyte was evaluated and confirmed following two rules: (1) the retention times matched the analytical standards with no more than 2% variation and (2) detected peaks had the correct precursor/product ions for both the quantitation and confirmation ions. The ratio of the quantitative precursor/product ion pair to the confirmation precursor/product ion pair in the sample did not vary by more than ± 30% from the ratio observed in calibration standards. To evaluate method accuracy and precision, six replicates of matrix blank, QCL, QCML, QCMH, and QCH were measured over different days. Three injections were made for each extract. As no endogenous amount of the target analyte was found in the matrix blank, accuracy was simply calculated as the calculated concentration divided by the spiking concentration. Precision was calculated as a percent relative standard deviation (RSD) of the calculated concentration. Because a comparable method for these analytes is not currently commercially available, external QCs were not performed by an outside lab for this study.

2.6. Optimization of sample preparation

The Oasis HLB SPE cartridge (Waters Corporation, Milford, Massachusetts, USA) is commonly used to extract neonicotinoids from water;^{[21, 62–64]} however, Gbylik-Sikorska et al. found that Strata X-CW cartridge could provide better recoveries for both neonicotinoids and their metabolites, such as imidacloprid urea and imidacloprid olefin, in honeybees and honey.^[65] In this study, both cartridges were tested, and acidic to basic sample media were considered for each cartridge to optimize SPE conditions. Accordingly, process efficiency was calculated and compared for each analyte under acidic, neutral, and basic conditions.

To optimize the sample preparation procedure, Iowa River water was collected and spiked with stock solution to obtain a matrix spike with analyte concentration of 10.42 ng/L. To further evaluate the influence of sample pH, we investigated its effect on the extraction process using formic acid (FA) or triethylamine (TEA) to adjust the water sample (pH = 7.4) to acidic (0.25% FA; pH = 2.5) or basic (0.33% TEA; pH = 11.7) conditions. Notably, this range is much wider than the pH values measured across the sample types analyzed in this study (pH 7.6 ± 0.5). Acidic and neutral conditions were tested for the HLB cartridge while acidic, neutral, and basic conditions were tested for Strata-X-CW. Water samples were then processed as previously described in section 2.4, except that 6 mL of acetonitrile/ ethyl acetate (8:2 v: v) was used as the elution solvent without adding any formic acid. For each condition, duplicates were prepared and extracted.

Next, the extracts were analyzed, and for each analyte, the averaged peak area was used to calculate the process efficiency (PE). PE is a comprehensive approach for evaluating analytical method development by comparing the response of the analyte in the matrix to the response in a pure standard solution. PE evaluates the method’s overall performance in real sample conditions by considering both matrix effects and analyte recovery after the extraction process. It is an important parameter to obtain when optimizing the extraction process. The formula for PE is as follows:^[66]

$$\text{PE}\left(\%\right)=\frac{\text{B}}{\text{A}}\times 100$$

where A represents the average peak area of the standard solution, and B corresponds to the peak area of an extract with the same concentration of the standard solution, prepared and extracted from a matrix spike. In LC-MS/MS, the most commonly observed matrix effect is ion suppression (i.e., a loss of ion intensity of the target analyte).^[67] A PE value below 100% is caused by matrix effects and loss of analyte during the extraction process, thus a higher process efficiency is more desirable.

2.7. Data Analysis

Water samples were analyzed following the analytical method developed and validated in this study. Raw data from the analytical run were integrated using SCIEX OS 3.3.1.43 software (ABSciex, Framingham, Massachusetts, US), and the peak area was used for quantification. A batch normally started with three injections from a medium range calibrator to equilibrate the system, one solvent injection to show any instrument background, one lab reagent blank with ISTD (LRB), 10–12 calibrators, a water rinse, at least one matrix spike (QCL, QCM, or QCH) every ten unknown samples, one Milli-Q water extract, and up to 60 water samples. We note that the unknown samples used for matrix spikes consisted of all matrices investigated herein: well water, surface water, and wastewater effluent. Because method validation was only performed with deep well water, we elected to use different matrices to ensure the method was valid for each type of water sample. The accuracy of target analytes in these matrix spikes was within the acceptable range (70–130%), suggesting that differences in the composition of well water, surface water, and wastewater effluent had little influence on the quantification results. In addition to these matrix spikes, one calibration standard was also checked every ten samples to ensure that the instrument response did not significantly change over time.

The detection frequency (DF) of each analyte was defined as the proportion of calculated concentrations >LOD. For samples in which no target analytes were observed >LOD, the entry was indicated with “ND” (no detect), and zero was used for statistics calculations. For those in which target analytes were between LOD and LOQ, half of the LOQ was used for statistics calculations.

3. Results and discussion

3.1. Method Development and Validation

The process efficiency (PE) of each analyte under varying SPE conditions is shown in Figure 1. The value for most analytes was between 30%−60% across the range of conditions and SPE cartridges considered. There were notable instances where the sample pH influenced SPE process efficiency across the various analytes. First, the PE of THX-U was close to 90% with Strata-X-CW after adjusting the water sample to be basic; however, the values for 5-OH-IMI, DIN-N-DES, and NIT-N-DES were less than 20% under this condition. Second, the PE of NIT was extremely low with Strata-X-CW for an acidic water sample. Accordingly, these two conditions were deemed not suitable for further method development. PE values were similar when comparing the remaining three conditions (HLB with acidic or neutral samples, Strata-X-CW with neutral samples) while Strata-X-CW with neutral water samples was slightly better than the other two conditions, especially for 5-OH-IMI, DIN-N-DES, and IMI-O. Thus, we elected to use the Strata-X-CW cartridge without adjusting sample pH for this method. A PE value below 100% is attributed to both matrix effects and analyte lost during extraction and in this study, it was mainly from the matrix effects. To solve this, isotope labeled internal standards, which were expected to have identical or similar behavior during extraction and ionization properties during instrument analysis, were used to compensate any matrix effect. This resulted in high accuracy and precision, which would be showed and discussed later in the method validation process.

Figure 1.

Process efficiency (%) of neonicotinoids and their metabolites in Iowa river water (10.42 ng/L) using two different solid phase extraction cartridges: HLB (Waters) and Strata-X-CW (Phenomenex) at different pH conditions: acidic (0.25% formic acid (v/v), neutral, basic (0.33% TEA).

We note that D-IMI was not detected in any of the aforementioned conditions. We subsequently discovered that introducing 0.25% formic acid to the elution solvent [acetonitrile/ ethyl acetate (8:2 v: v)] improved recovery for D-IMI, but its recovery in samples was not consistently reproducible. As such, we have elected not to include the results associated with D-IMI as part of this study, and we continue to investigate how factors such as pH, ionic composition, or other sample conditions affect the detection of D-IMI.

Linearity, LOD, LOQ, precision, and accuracy (relative recovery) were evaluated for method validation, and the results are summarized in Table 1. The calibration curves showed good linearity (r²>0.99) for all analytes over the range from LOD to 10 ng/mL, except for DIN-N-DES, for which the linear range was LOD to 5 ng/mL. The high end of the linear detection range was ultimately extended to 25 ng/mL for CLO and IMI because their concentrations were frequently found to be higher than 10 ng/mL during analysis of authentic water samples. The LOD and LOQ of this method were determined using well water spiked with a very low concentration of analytes (QCL, 0.42 ng/L). LODs are between 0.02–0.15 ng/mL in the extract, which corresponds to 0.16–1.21 ng/L in water samples. LOQs are between 0.06–0.48 ng/mL in the extract, corresponding to 0.54–4.03 ng/L in water samples. DIN, NIT, and their metabolites DIN-N-DES and NIT-N-DES, respectively, have a higher LOD and LOQ values compared to the rest of analytes.

Table 1.

Method validation results for neonicotinoids and their metabolites spiked to well water samples at four different concentrations (target concentration: QCL (QC Low) = 0.42 ng/L, QCML (QC Medium Low) = 4.17 ng/L, QCMH (QC Medium High) = 10.42 ng/L, and QCH (QC High) = 25 ng/L; n = 18).

			QCL			QCML			QCMH			QCH
Compound	LOD ng/L	LOQ ng/L	Accuracy (%)		Precision (%)	Accuracy (%)		Precision (%)	Accuracy (%)		Precision (%)	Accuracy (%)		Precision (%)
			Ave	SD	RSD	Ave	SD	RSD	Ave	SD	RSD	Ave	SD	RSD
ACE	0.18	0.60	109.7	14.4	13.2	94.4	7.0	7.4	99.1	8.1	8.2	97.2	5.2	5.3
CLO	0.27	0.91	98.3	22.1	22.3	91.2	6.6	7.3	95.4	13.3	14.0	91.0	4.4	4.9
IMI	0.27	0.92	130.8	22.0	16.8	91.5	7.5	8.2	94.7	8.3	8.8	96.4	4.1	4.3
THC	0.16	0.54	94.2	13.1	13.8	87.2	4.5	5.2	90.5	5.3	5.9	91.2	2.4	2.6
THX	0.35	1.17	132.7	28.5	21.3	92.6	8.0	8.6	93.7	8.8	9.4	92.8	7.2	7.8
DIN	0.83	2.06	150.7	49.3	32.8	89.7	9.1	10.1	97.6	8.0	8.2	94.7	4.5	4.7
NIT	1.21	4.03	128.9	96.8	74.9	92.7	12.7	13.7	97.6	9.1	9.4	104.9	6.8	6.4
FLU	0.24	0.81	99.9	19.3	19.5	92.8	11.2	12.0	92.7	12.2	13.1	95.4	7.5	7.9
FLO	0.22	0.75	123.3	17.8	14.5	96.0	8.8	9.1	97.9	9.1	9.3	97.0	5.6	5.8
SUL	0.24	0.79	112.6	19.2	16.9	105.8	8.4	8.0	109.8	11.3	10.3	107.6	7.8	7.2
IMZ	0.41	1.37	101.7	33.1	32.3	90.5	7.6	8.3	94.8	6.5	6.8	96.1	5.4	5.6
ACE-N-DES	0.22	0.75	107.8	17.8	16.7	90.5	6.3	7.0	94.5	6.4	6.8	96.5	5.6	5.9
5-OH-IMI	0.38	1.27	123.9	30.5	24.7	95.2	12.4	13.1	97.7	9.2	9.4	95.9	5.9	6.1
THX-N-DES	0.17	0.56	100.2	13.1	13.3	83.1	6.6	7.9	89.0	7.6	8.6	87.4	5.7	6.5
CLO-U	0.36	1.19	139.8	28.6	20.4	89.8	6.6	7.4	95.2	7.7	8.1	99.9	4.8	4.8
THX-U	0.25	0.85	113.5	20.4	17.9	85.2	16.0	18.7	91.1	14.1	15.5	90.0	13.1	14.5
IMI-U	0.19	0.63	87.4	15.0	17.2	74.7	16.6	22.3	79.3	10.9	13.8	84.8	8.3	9.8
THC-A	0.26	0.88	121.4	21.1	17.4	91.1	3.9	4.3	92.6	6.9	7.4	94.6	5.7	6.1
DIN-N-DES	0.83	1.21	71.7	29.3	40.6	63.7	12.4	19.4	101.8	8.8	8.7	87.3	5.4	6.2
NIT-N-DES	0.83	2.78	196.9	66.8	33.9	88.8	13.7	15.4	72.1	12.5	17.3	85.0	11.2	13.2
CLO-N-DES	0.41	1.36	117.6	32.6	27.6	90.3	10.7	11.9	97.9	13.3	13.6	96.1	10.3	10.7
SUL-X	0.24	0.79	92.2	19.0	20.5	78.9	12.4	15.7	84.8	14.9	17.5	80.0	9.4	11.8
IMI-O	0.37	1.23	102.0	29.3	28.8	101.3	9.0	8.9	110.1	12.5	11.3	119.1	14.0	11.7

The accuracy of QCML (4.17 ng/L) for neonicotinoid parents ranged from 87.2 to 94.4%, from 74.7 to 101.3% for the NEO metabolites, and 92.8 to 105.8% for the 3 related insecticides. The QCML accuracy for DIN-N-DES, which was excluded from the final method, was 63.7%, which is outside of our acceptable range of 70 to 130%. Precision values were below 15% for parents and below 20% for metabolites at this concentration level, except for IMI-U, which was 22.3%. At the medium high concentration level (QCMH, 10.42 ng/L), accuracy ranged from 90.5–109.8% and 72.1–110.1% for parents and metabolites, respectively, with a precision less than 20% for all analytes. For QCH (25 ng/L), the accuracy ranged from 91.0–107.6% and 80–119.1% for parents and metabolites, respectively, with a precision less than 15% for all analytes. Even at the lowest concentration level we tested (QCL, 0.42 ng/L), most of the analytes showed acceptable recovery (i.e., between 70 and 130%) except for four analytes: THX, DIN, CLO-U, and NIT-N-DES. We note that because LOD and LOQ values can vary significantly when applying different calculation methods,^[68–71] these values are best rationalized experimentally using accuracy and precision values. For example, we calculated an LOQ of 0.6 ng/L for ACE. This value was confirmed experimentally by preparing the QCL at a concentration level of 0.42 ng/L and demonstrating acceptable accuracy and precision (109.7 and 13.17%, respectively) for this sample.

All method parameters produced herein are comparable to or better than other analytical methods developed for the study of NEOs. For example, one prior study^[28] used SPE to process 1 L of water for measurement of six neonicotinoids, reporting method detection limits (MDLs) ranging from 3.6 to 6.2 ng/L. Another study reported a method detection limit (MDL) of 0.03–0.2 ng/L for seven neonicotinoid parent compounds after enriching 1 L of water using SPE, and the QC low samples used in method validation were prepared at 25 ng/L.^[21] Another study evaluated ten neonicotinoids and two metabolites using 1 L water and reported 0.01–0.05 ng/L as the LOQ; however, the QC samples were prepared at a much higher concentration level, 10 ng/L.^[63] The method developed herein was able to achieve these performance measures for a broader number of NEOs and their metabolites while also using a much smaller sample volume of 12 mL, highlighting the novelty and value of our approach.

3.2. Analytes in Iowa well water

In this study, eight wells from three Counties in Iowa (Scott County, Johnson, and O’Brien County) were sampled six to eight times from April to November in 2023. Results of the analysis for private well water samples from Scott (S), Johnson (J) and O’Brien (O) counties are shown in Figure 2 and Table S7 based on their DF (Figure 2A) and cumulative neonicotinoid concentration (on a mass basis; ng/L) measured in each water sample (Figure 2B). Three neonicotinoids, CLO, IMI, and DIN, and the butenolide, FLU, were detected. For comparison, Thompson et al. previously evaluated NEOs in 40 untreated well water samples. CLO and IMI were commonly found with DFs of 68% and 43%, respectively.^[21] THX was also detected in one sample in that study.^[21]

Figure 2.

Monitoring of NEOs in eight wells from April to October: (A) Detection frequencies of four neonicotinoids (S1: n = 8, S2: n = 8, S3: n=6, J1: n = 7, J2: n =7, O1: n =7, O2: n =6, O3: n = 6) and (B) seasonal variations for the total concentrations from each site; S: Scott County; J: Johnson County; O: O’Brien County.

As shown in Figure 2A, CLO was not detected in the three private wells from Scott County (S1–S3) but detected frequently in the wells from Johnson (J1–2) and O’Brien Counties (O1–3). The CLO DF was 57.1% for both sites in Johnson and 52.9%, 83.3% and 100% for three well sites tested in O’Brien County. The median CLO concentration was 0.5 ng/L for both sites in Johnson County (Table S7), which is consistent with private well water concentrations for CLO previously reported by Thompson et al. (0.4 ng/L).^[21] The median level of CLO ranged from <LOD to 3.8 ng/L for the three well in O’Brien County, and the maximum concentration levels (3.88 – 6.19 ng/L) were comparable to results reported previously for Northwest Iowa^[20], but considerably lower than those reported by Thompson et al. in other agriculturally intensive areas in Iowa (391.7 ng/L).^[21] These differences are likely due to distinct geology, well depth, or aquifer types in each location, which have been shown to be significant factors associated with variations in NEO concentrations across the state.^[20]

IMI was detected in all well samples, except for two wells in O’Brien County (Figure 2A). A 100% DF was found in J1, with a median concentration of 3.3 ng/L. IMI was also detected once in S2 in June with a concentration of 27.2 ng/L, which was higher than the maximum detected concentration (6.7 ng/L) in our prior well study.^[21] DIN was also frequently detected at site J1 with a DF of 85.7% (Table S7) and median detect concentration of 3.2 ng/L. FLU was detected twice in the same site in O’Brien County (O2). This is notable, as before this study, FLU was never previously reported in Iowa water.

Overall, total mass concentration summed for all NEO was relatively low in these private wells (all less than 30 ng/L; see Figure 2B). Moreover, we observed no significant seasonal variation in total NEOs across our study period. NEOs were more prevalent in shallow wells (in O’Brien and Johnson County relative to deeper wells in Scott County). Generally, more NEOs were found in private wells in agriculturally intense areas (O’Brien County) than in more urban areas like Johnson and Scott County. This is consistent with findings reported in 2021 that showed median concentrations of CLO (p = 0.01) and THX (p = 0.01), total NEO concentration (p = 0.01), and number of detects per well (p = 0.01) were each significantly higher in Northwest Iowa (O’Brien County) compared to the rest of state.^[20] In that study, the DF for Northwest Iowa was 56% compared to 13% for Southeast Iowa, where Johnson and Scott counties are located.^[20] Average total NEO concentrations per well were also 6 times higher in Northwest Iowa (4.4. ng/L compared to 0.7 ng/L).^[20] Overall, these analyses support prior studies suggesting relatively widespread occurrence of NEOs in well water in Iowa, albeit at low levels. The near-consistent occurrence of NEOs in private well water from April to October, particularly in shallow wells in agricultural settings, is consistent with their prevalence in shallow aquifers in such regions.

3.3. Analytes in Iowa surface water

Iowa surface waters were sampled eight times across 2023 from 5 locations - Mississippi River (MS), Burr Oak Creek (BO), Little Floyd River (LF), Pilot Creek (PC), and Cardinal Creek (CC). Five NEO parents, ACE, CLO, DIN, IMI, THX, and the butenolide, FLU, as well as six metabolite compounds, ACE-N-DES, CLO-N-DES, IMI-U, IMI-O, THX-U, and CLO-U, were detected in at least one location. To the best of the authors knowledge, this is the first report of the metabolites CLO-N-DES, THX-U, and CLO-U, in US water samples.

The three primary NEOs used in the Midwest, CLO, IMI, and THX,^[20] were frequently detected at all five sites, as shown in Figure 3. For CLO, the DF ranged between 50% to 90%. The median concentration of CLO was between 0.2 and 12.9 ng/L, with the maximum detectable level of 375.2 ng/L (Table S8) found in CC. The DF of IMI was higher than that observed for CLO, ranging between 87.5% to 100% across all locations. The median concentration of IMI was 0–2.5 ng/L, except in CC where the median concentration was an order of magnitude higher at 36.9 ng/L. THX was frequently detected at all locations as well (Figure 3), with high DF (87.5%) in BO (median = 12.9 ng/L, maximum = 214.2 ng/L) and PC (median = 11.7 ng/L, maximum = 14.9 ng/L). The DF of THX in MS was 25%, which was lower than the frequency reported from the same site in 2013 (50%),^[28] while the maximum concentration detected increased to 11.1 ng/L (Table S8) from 5.6 g/L.^[28]

Figure 3.

Detection frequency of six NEOs and six metabolites in surface water from March to October in 2023 from five locations (n = 8 for each): MS (Mississippi River), BO (Burro Oak Creek), LF (Little Floyd), PC (Pilot Creek), and CC (Cardinal Creek).

Our findings with CLO and IMI are noteworthy both for their frequency and levels of detection. In terms of DF, a USGS study reported that CLO was most frequently detected with a DF of 75% in seventy-nine water samples from a network of nine sites across Iowa in 2013, while the overall DF of IMI was only 23%.^[28] The increase of IMI DF to 97% is consistent with increasing use of IMI in Iowa and the Midwest over the past ten years. For example, the same study monitored the Mississippi River in 2013 from early May to late October and IMI was not detect,^[28] while our study showed the DF of IMI was close to 90% for the same water body in 2023. The levels detected in the Mississippi River are also notable considering that the samples from late April and May of 2023 were collected when the River was at flood stage. Based on the volumetric flow (215,380 cubic feet per second or CFS) measured nearby at Lock and Dam 14 at Le Claire, IA, on May 9, 2023 (i.e., the date we measured a maximum concentration of CLO of 95.9 ng/L), we estimate a CLO mass load of just over 50 kg/d (about 110 lbs./d) for the Mississippi River.

Chronic toxicity benchmarks for aquatic invertebrates set by the US EPA for IMI, CLO, and THX are 10 ng/L, 50 ng/L, and 740 ng/L, respectively (Table S1).^[43] ALBs assess ecological risk and are used to protect aquatic organisms, including fish, invertebrates, and plants from harmful effects. The ALBs serve to protect freshwater species by considering the lethal and sub-lethal impacts (such as immobility, slower growth, and reduce reproduction) caused by pesticide residues in water.^[43] In terms of levels of detection, the maximum detectable level of CLO in CC (375.2 ng/L) well exceeds the chronic toxicity ALB for freshwater invertebrates of 50 ng/L (Table S1), whereas the median concentration of IMI in CC (44 ng/L) was also over 4-fold higher than its chronic toxicity benchmark for freshwater invertebrates at 10 ng/L.^[43] The frequent detection of NEOs at levels above the chronic toxicity benchmarks in CC suggest the insecticides could pose a risk to non-target aquatic invertebrates due to long-term chronic exposure. In contrast, for the prior USGS study on NEO occurrence in Iowa surface waters, the median levels reported were below the chronic toxicity benchmarks for invertebrates at all sites for CLO and IMI.^[28] The observed differences between the USGS study and the reported results are likely due to variations in the sampled sites. The researchers are uncertain if this discrepancy indicates a broader increase in NEO concentrations in surface waters in Iowa over time.

DIN and FLU were also detected in samples from at least one surface water site. DIN was detected most frequently in CC (DF = 100%, median = 36.9 ng/L, maximum = 191.4 ng/L), followed by LF (87.5%, median = 2.5 ng/L, maximum = 31.8 ng/L), and BO (25%, median = 0.0 ng/L, maximum = 1.0 ng/L). This represents far greater frequency of detection for DIN compared to prior monitoring efforts of Iowa streams; for example, the prior USGS study only detected DIN once but at levels below their MDL.^[28] FLU was also detected twice in CC with a maximum concentration of 1.0 ng/L. ACE and its metabolite ACE-N-DES were detected three times in CC but at concentrations lower than their LOQ. This is likely because ACE is used less in Iowa compared to other NEOs.^{[28, 30]}

CLO-N-DES and CLO-U are two metabolites of CLO.^[72] CLO-N-DES was only detected in BO (DF = 12.5 %, median: ND, maximum = 6.2 ng/L) while CLO-U was detected in BO (DF = 62.5%, median = 6.8 ng/L, maximum = 31.7 ng/L), LF (DF = 12.5%, median: ND, maximum = 2.7 ng/L), and CC (DF = 22.2%, median: ND, maximum = 6.0 ng/L) (see Table S8). Overall, the DF was 3% and 21% and the maximum concentration was 6.2 ng/L and 31.7 ng/L for CLO-N-DES and CLO-U, respectively. The median was below LOD for both metabolites. A previous study reported both CLO-N-DES and CLO-U in drinking water samples (n = 884) from 32 provinces throughout China in June 2019 with DF of 3.2% and 45% for CLO-N-DES (medium: ND, maximum = 2.25 ng/L) and CLO-U (medium: ND, maximum = 15.1 ng/L), respectively.^[72] Our results are consistent with these previous reported DFs while the maximum concentrations are slightly higher due to the maximum concentrations observed at BO, which is likely impacted by wastewater via treatment lagoon discharge. In addition, CLO, CLO-N-DES, and CLO-U are also three major metabolites of THX,^[73] as THX can be degraded through hydrolysis, photolysis, and microbial biodegradation.^{[50, 51, 74]} This might explain why CLO-N-DES and CLO-U were not detected in MS even though CLO was highly detected (DF = 87.5%).

Besides CLO, CLO-N-DES, and CLO-U, THX can also undergo N-demethylation to form THX-N-DES or convert to THX-U via a reduction reaction.^[73] THX-N-DES was not detected in this study while THX-U was detected in BO (DF = 87.5%, median = 91.7 ng/L, maximum = 416.7 ng/L), LF (DF = 12.5%, median: ND, maximum = 0.9 ng/L), and PC (DF = 25%, median: ND, maximum = 2.6 ng/L) (see Table S8). Incidentally, THX-U was the most frequently detected NEO metabolite in China drinking water as the DF = 76% and median = 0.94 ng/L.^[72]

For the metabolites of IMI, IMI-U and IMI-O were frequently found in at least one sampling site. IMI-U was detected most frequently in CC (62.5%, median = 1.3 ng/L, maximum = 6.8 ng/L), followed by BO (25%, median = 0 ng/L, maximum = 1.8 ng/L), and it was also detected once in LF and PC (see Figure 3). IMI-O was also detected once in CC, but the concentration was below LOQ. Overall, the DF was 24% and 3% for IMI-U and IMI-O, respectively.

Although IMI-U and IMI-O could be formed in soil and water from IMI as its degrades via hydrolysis, photolysis or ultraviolet irradiation, and microbial biodegradation, the occurrence of IMI metabolites is seldom documented.^{[50, 51, 56, 75–80]} Wan et al.^[62] evaluated five degradates of IMI in source, treated, and tap water in central China and detected IMI, IMI-U, and IMI-O with 100% frequency in source and treated water samples. They also reported between 0.78–1.51 ng/L of IMI-U in the surface source water in July 2019.^[62] The DF of the two IMI metabolites reported in China are significantly higher than the numbers in our study, which might indicate heavier use of IMI in China. Klarich Wong et. al.^[56] detected between 0.08–0.66 ng/L of IMI-U in raw source water collected from the University of Iowa and Iowa City drinking water treatment plants, about ten-fold lower than the concentrations found in CC in this study (0.3–6.8 ng/L). Because metabolites of certain NEOs, including IMI, are known to exhibit greater mammalian toxicity, the frequency and level of IMI metabolites detected herein suggest the need to monitor for the occurrence of IMI metabolites in water resources where NEOs are heavily used.

3.4. Neonicotinoid Prevalence in Effluent Impacted Streams

It is notable that all NEOs and metabolites, except THX and its metabolites, were more frequently detected and present at higher concentration levels in CC, an effluent impacted stream, than at any other site, including several creeks (e.g., PC and LF) located in more agriculturally intense areas. Figure 4A shows the total summed concentration of each parent NEO and its metabolites at CC over the sampling duration. The Total CLO (= CLO + CLO-N-DES) concentration peaked in August (375.2 ng/L) and October (196.2 ng/L) of 2023 and remained high in follow-up sampling in March of 2024. As shown in Figure 4A, values of Total CLO were frequently above the US EPA ALB for freshwater invertebrates in CC. Similarly, there was little to no temporal variation in Total IMI (= IMI + IMI-U + IMI-O) concentration across all samples at CC, with all Total IMI levels also above the US ALB for freshwater invertebrates across the one-year sampling period.

Figure 4.

Monitoring of NEOs in CC from March 2023 to March 2024: (A) seasonal variations for concentrations of each NEO and its metabolites. The aquatic life chronic benchmarks for freshwater invertebrates are shown (10 and 50 ng/L for imidacloprid and clothianidin, respectively). (B) Map of CC in the city of Newton, Jasper County in Iowa. Samples were collected from three sites (upstream, wastewater effluent, downstream) in March 2024. (C) The concentration of four NEOs (ng/L) found in the upstream, wastewater effluent, and downstream in CC.

To better understand the high NEO concentrations in CC, we also conducted follow-up sampling in March 2024 upstream of the wastewater effluent discharge and pure wastewater effluent for comparison to measurements of our typically collected downstream samples (Figure 4B). As shown in Figure 4C, only 1.4 ng/L of IMI was found in upstream water, whereas 29.6 ng/L of IMI was found in the wastewater effluent and 31.3 ng/L in the downstream. In addition, 198.9 ng/L of CLO was found in the wastewater effluent and 162.8 ng/L in the downstream. Assuming an effluent discharge of 3 MGD, which is typical for dry weather operations at the Newton Sewage Treatment Plant, this effluent CLO concentration corresponds to an effluent loading of approximately 2 g/d.

Collectively, our results indicate that the high NEO concentrations found in CC are attributable to effluent discharge from the wastewater treatment in Newton. NEOs are known to be only sparingly removed during conventional wastewater treatment processes.^{[11, 81, 82]} Although NEOs in Iowa likely come from a variety of sources including agriculture^[20], urban land use^[27], and domestic sources like pets^[64], wastewater effluent may be a significant source of NEOs and their metabolites in the environment,^{[11, 81–85]} especially during dry periods where wastewater effluent may represent a large fraction of flow in some streams. Indeed, others have also suggested potential risks to the ecosystem from treated wastewater discharged into effluent-dominated, receiving surface waters.^[82] Further investigation is needed to better understand the sources of NEOs within the service area of the wastewater treatment plant (i.e., the sewershed) and the persistence of these compounds in effluent post-treatment and their potential impact on the environment.

BO represents another wastewater effluent impacted site in our study. A notable observation at BO was the significant seasonal variation of THX, as well as its metabolite. As shown in Fig 5A, the amount of THX in BO ranged from 4.4 to 6.2 ng/L from March to June and then peaked at 214.2 ng/L from June to August. Concentrations then decreased to 12.9–16.4 ng/L in September and October. As a transformation product of THX,^[28] CLO did not show this seasonal variation and this indicates that most of CLO occurrence likely came from its direct use; however, similar trends in concentration over time were observed for the other two THX metabolites, THX -U and CLO-U, in BO (Figure 5A). The DF of THX-U and CLO-U were 87.5% and 62.5% at this site, and the concentrations of both species increased to their maximum values in August (416.7 ng/L and 31.7 ng/L for THX-U and CLO-U, respectively). In September, the concentration of THX-U remained high (408.3 ng/L) while the concentration of THX and CLO-U started to decline. CLO-N-DES, a common metabolite of CLO, was detected once (6.2 ng/L) at BO in August, when the concentration of CLO was low (0.5 ng/L), indicating it may be derived from THX instead of CLO.

Figure 5.

Seasonal variation of (A) thiamethoxam and its four metabolites (thiamethoxam metabolite CGA 355190, thiamethoxam metabolite CGA 353968, clothianidin, and clothianidin-n-desmethyl) and (B) the total concentration of thiamethoxam and its metabolites as well as clothianidin and its metabolites from BO (Burr Oak Creek). The aquatic life chronic benchmark for freshwater invertebrates is 50 and 740 ng/L for clothianidin and thiamethoxam, respectively.

Total CLO (= CLO + CLO-N-DES) and Total THX (= THX + THX-U + CLO-U) over time in BO are shown in Figure 5B. Also included is the USGS gage height at BO illustrating potential changes in stream flow in response to rainfall events. Total THX concentration increased relatively steadily over time and does not appear to respond to changes in flow in BO induced by rainfall events. The peak concentration in August is also well after Spring planting season (typically late April or early May in Northwest Iowa), and Total THX concentrations remain elevated even after harvest (into November). Thus, we speculate that the temporal trends in Total THX may reflect the changing levels of these species present in the lagoon effluent discharged into BO upstream of our sampling site. Regardless of the source, the August peak in Total THX approaches the ALB for chronic toxicity to freshwater invertebrates for THX (740 ng/L). Because of the considerable contribution of THX metabolite concentrations to Total THX levels, it may be important to expand monitoring for THX-U and CLO-U and include these species when assessing the potential ecological and human health impacts associated with THX use.

4. Conclusions

The method developed in this study demonstrates significant improvements in sensitivity and efficiency compared to previous methods for the analysis of NEOs and their metabolites in water samples. The application of this method to well water samples from Iowa revealed the widespread presence of NEOs, including CLO, IMI, and DIN, with notable detection of the butenolide FLU for the first time in Iowa well water. The analysis showed higher concentrations of NEOs in wells from agriculturally intensive areas compared to urban areas, aligning with previous findings. In surface water samples, five primary NEOs and six metabolites were frequently detected across various locations. The study highlighted an increased frequency and concentration of IMI compared to previous studies, indicating its growing use in the region. We also observed frequent detections of CLO and IMI in the Mississippi River, at times corresponding to very large NEO mass loadings during times of high flow. Notably, the detection of NEOs and their metabolites in an effluent-impacted streams including CC and BO underscored the potential role of wastewater discharge in NEO contamination of surface waters, especially for smaller creeks and streams. For effluent impacted streams, we anticipate that metabolites will be particularly important components of certain types of wastewater discharge; for example, trends of Total THX observed at BO may reflect formation of CLO, CLO-N-DES, and CLO-U, the major metabolites of THX, during biological degradation in the wastewater treatment lagoon that feeds the creek. Finally for the duration of our study, Iowa was in the midst of a multi-year drought. We contend that under such conditions of relatively low stream flows, more attention should be paid to municipal wastewater effluent as a source of neonicotinoids, and more effort is needed to understand sources of neonicotinoids discharged by municipal wastewater facilities. Overall, this study highlights the prevalence and persistence of NEOs in Iowa water sources and raises concerns about their potential ecological impacts due to persistently elevated levels above EPA ALBs.

The developed method provides a reliable tool for future monitoring and risk assessment of NEO contamination in environmental water samples, although there remain opportunities for further improvement. The method was unable to reliably quantify potentially important metabolites including desnitro-imidacloprid. Thus, future analytical method development should aim to expand the number of metabolites, especially those of most widely used CLO, IMI, and THX, for monitoring in water resources. Moreover, we acknowledge that the matrices investigated herein may not fully represent the complexity of all water matrices, especially in areas with varying levels of agricultural runoff or other sources of contamination. Indeed, seasonal variations and localized factors could influence the occurrence and concentration of NEOs, and the method will require additional spatial and temporal sampling to ensure its broad applicability. Similarly, our study focused solely on water sources in Iowa, which may limit the generalizability of the findings to other regions with different bulk water quality (e.g., alkalinity, ionic content, organic matter), agricultural practices, pesticide usage, and environmental conditions. Further research is recommended to validate the method’s applicability across diverse water systems and geographical areas.